metal accumulation in a constructed surface flow...

UNIVERSITEIT

GENT

FACULTY OF BIOSCIENCE ENGINEERING

CENTRE FOR ENVIRONMENTAL SANITATION ____________________

Academic Year 2004 – 2005

METAL ACCUMULATION IN A CONSTRUCTED SURFACE FLOW WETLAND

ACCUMULATIE VAN METALEN IN EEN

VLOEIRIETVELD

Maja ŠIMPRAGA

Promoter: Prof. dr. ir. Filip Tack Co-promoter: ir. Els Lesage

Laboratory of Analytical Chemistry and Applied Ecochemistry

Coupure Links 653 9000 Gent

Master Thesis to obtain the degree of M.Sc. in ENVIRONMENTAL SANITATION

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COPYRIGHT

The author, the promoter and the co-promoter give the permission to use this dissertation for

consultation and to copy parts for personal use. Every other use is subjected to the copyright

laws. The source must be extensively specified, when using results from this dissertation.

Ghent, Belgium

July, 2005.

Promoter _________________________

Prof. Dr. ir. Filip Tack

Co-promoter ______________________

Ir. Els Lesage

Author____________________________

Maja Šimpraga, dipl.ing

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Mami i Tati

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‘Here again we are reminded that in nature nothing exists alone’

Rachel Carson, ‘Silent spring’

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ACKNOWLEDGEMENTS

Hereby, I would like to thank all professors and colleagues that helped and assisted me during

this research.

First of all I would like to express my gratitude to prof. dr. ir. Filip Tack, who encouraged and

gave me the possibility to execute this dissertation.

I would like to express my sincere gratitude to ir. Els Lesage, who patiently initiated and

guided me throughout my work. Thank you for all these hours spent working together.

I should also like to acknowledge my debt to the Center of Environmental Sanitation, prof. dr.

Marc Van den Heede, Helga, Veerle, Isabelle, and Sylvie for mutual understanding, together

with prof. dr. ir. Marc Verloo, prof. dr. Niels De Pauw and prof. dr. Jan Pieters, that provided

me with a lot of new and up-to-date information.

My special thanks goes to Tom Kiekens for tirelessly editing the manuscript and providing

me with some useful ideas. Tommi, thanks for a big support and a chance to go forward.

I cannot close these acknowledgements without giving special gratitude to my family, brother

Saša, and sister Sanja for the support they gave me throughout these years of study. Also, I

must not forget Baka, my grandma, for the special support.

Ria, Kathy, Steven, Martin, Yasin, and Saroj thank you for nice working moments.

May all those who have played a part in this dissertation be included in this expression of

gratitude.

Maja

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SUMMARY

Accumulation of heavy metals was studied in a surface flow constructed wetland (CW) that

has been treating domestic wastewater since 1989. The CW was composed of a presettlement

tank, 9 surface flow channels planted with Phragmites australis, and a ditch in which the

effluent was collected before discharge into the surface water. Sediment, aboveground plant

parts (leaves, stems, leaf sheaths, and panicles), and wastewater were collected and analysed

on heavy metals. The majority of metals were retained in the sediment of the CW, whereas

metal accumulation in aboveground plant parts was very low. Total metal concentrations in

the sediment did not exceed soil sanitation criteria, but were elevated compared to

background values. The pollution level of the sediment was low to moderate. Metal mobility

was estimated by means of the exchangeable metal fraction and the SEM/AVS ratio. The

exchangeable metal fraction was generally low, except for Cd, Mn, and Zn of which 15 – 39,

12 – 14, and 6 – 10 % of the total amount is in an exchangeable state, respectively. SEM/AVS

ratios were lower than 1 indicating that Cd, Cu, Ni, Pb, and Zn were not potentially available.

However, the use of the SEM/AVS ratio as a single measure of potential availability is not

suggested. Concentration levels in the wastewater were generally low and high removal

efficiencies were seen for Al, Cu, and Zn (28 – 63, 63 – 78, and 57 – 62 %, respectively). Mn

appeared to be released from the CW at this stage of operation.

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SAMENVATTING

Deze studie onderzocht de accumulatie van zware metalen in een vloeirietveld die

huishoudelijk afvalwater behandelt sinds 1989. Het waterzuiveringssysteem bestaat uit een

voorbezinktank, 9 vloeigoten beplant met Phragmites australis, en een goot waarin het

gezuiverde effluent verzameld wordt voor lozing in het oppervlaktewater. Sediment,

bovengrondse plantendelen (bladeren, bladstelen, stengels, en pluimen), en afvalwater werden

bemonsterd en geanalyseerd op zware metalen. Zware metalen accumuleren voornamelijk in

het sediment en de bijdrage van de bovengrondse vegetatie aan de metaalaccumulatie wordt

als verwaarloosbaar beschouwd. De totale metaal concentraties in het sediment overschreden

de bodemsaneringsnormen niet maar waren evenwel verhoogd in vergelijking met de

achtergrondwaarden. De pollutiegraad van het sediment was laag tot matig. Metaal mobiliteit

werd geschat door middel van de uitwisselbare metaalgehalten en de SEM/AVS verhouding.

De uitwisselbare metaal fractie was laag behalve voor Cd, Mn, en Zn waarvan respectievelijk

15 – 39, 12 – 14, en 6 – 10 % van het totaal gehalte uitwisselbaar was. De SEM/AVS

verhouding was kleiner dan 1 en toont aan dat Cd, Cu, Ni, Pb, en Zn niet potentieel

beschikbaar zijn. Er wordt aangeraden om de SEM/AVS verhouding niet als enige parameter

te beschouwen bij het inschatten van metaal mobiliteit. Concentraties van zware metalen in

het afvalwater waren laag en hoge verwijderingsefficiënties werden genoteerd voor Al, Cu en

Zn (respectievelijk 28 – 63, 63 – 78, en 57 – 62 %). Mn concentraties in het effluent waren

hoger dan in het effluent en wijzen op vrijstelling van Mn in deze fase van operatie.

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TABLE OF CONTENTS

1. INTRODUCTION............................................................................................................... 1

2. LITERATURE REVIEW................................................................................................... 3

2.1. NATURAL WASTEWATER TREATMENT SYSTEMS AS AN ALTERNATIVE TO CONVENTIONAL WASTEWATER TREATMENT SYSTEMS .......................................................... 3

2.1.1. Aquatic wastewater treatment systems................................................................................................. 4

2.1.2. Terrestrial wastewater treatment systems............................................................................................ 5

2.1.3. Constructed wetlands........................................................................................................................... 6

2.2. HEAVY METAL ACCUMULATION IN CONSTRUCTED WETLANDS............................. 15

2.2.1. Removal processes of heavy metals in surface flow constructed wetlands ........................................ 16

2.2.2. Factors influencing metal mobility .................................................................................................... 17

2.2.3. Estimating metal mobility .................................................................................................................. 20

2.3. BELGIAN LEGISLATION FRAMEWORK ................................................................................ 22

2.3.1. Soil remediation criteria (Vlarebo, 1996).......................................................................................... 22

2.3.2. Surface water quality criteria ............................................................................................................ 23

3. OBJECTIVE OF THE STUDY ....................................................................................... 24

4. MATERIALS AND METHODS...................................................................................... 25

4.1. STUDY SITE ....................................................................................................................................... 25

4.2. SAMPLING.......................................................................................................................................... 27

4.2.1. Wastewater......................................................................................................................................... 27

4.2.2. Aboveground Phragmites australis biomass...................................................................................... 28

4.2.3. Sediment ............................................................................................................................................. 29

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4.3. ANALYTICAL PROCEDURES....................................................................................................... 30

4.3.1. Wastewater......................................................................................................................................... 30

4.3.2. Phragmites australis .......................................................................................................................... 31

4.3.3. Sediment ............................................................................................................................................. 31

4.4. MASS BALANCE .............................................................................................................................. 35

4.4.1. Metal mass removed from wastewater ............................................................................................... 35

4.4.2. Metal mass accumulated in aboveground Phragmites biomass......................................................... 36

4.4.3. Metal mass accumulated in the sediment ........................................................................................... 37

4.5. DETECTION LIMITS........................................................................................................................ 37

5. RESULTS........................................................................................................................... 38

5.1. WASTEWATER ................................................................................................................................. 38

5.1.1. pH and electrical conductivity (EC)................................................................................................... 38

5.1.2. Dissolved heavy metal concentrations ............................................................................................... 39

5.1.3. Removal efficiency of the SF wetland ................................................................................................ 42

5.1.4. Metal mass removed from the wastewater ......................................................................................... 43

5.2. PHRAGMITES AUSTRALIS BIOMASS .......................................................................................... 44

5.2.1. Metal concentrations in Phragmites australis biomass ..................................................................... 44

5.2.2. Metal mass accumulated in aboveground Phragmites biomass......................................................... 46

5.3. SEDIMENT .......................................................................................................................................... 48

5.3.1. Sediment characteristics .................................................................................................................... 48

5.3.2. Metal concentrations in the sediment................................................................................................. 51

5.3.3. SEM/AVS ratio ................................................................................................................................... 56

5.3.4. Metal mass accumulated in the sediment ........................................................................................... 58

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6. DISCUSSION .................................................................................................................... 59

6.1. METAL ACCUMULATION IN THE CONSTRUCTED WETLAND...................................... 59

6.1.1. Relative importance of sediment and Phragmites biomass in total metal accumulation in the SF

wetland......................................................................................................................................................... 59

6.1.2. Pollution level of the sediment ........................................................................................................... 60

6.1.3. Metal mobility in the sediment ........................................................................................................... 61

6.2. REMOVAL EFFICIENCY OF THE SURFACE FLOW WETLAND ....................................... 63

7. CONCLUSIONS................................................................................................................ 66

REFERENCES...................................................................................................................... 68

APPENDICES ....................................................................................................................... 75

GLOSSARY........................................................................................................................... 82

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ACRONYMS AND ABBREVIATIONS

AMINAL Flemish Environmental Administration (Administratie Milieu, Natuur, Land

en Waterbeheer van de Vlaamse overheid)

BOD Biological Oxygen Demand

CEC Cation Exchange Capacity

COD Chemical Oxygen Demand

CW Constructed Wetland

EMIS Energy and Environment Informational System for Flemish society (Energie

en Milieu Informatiesysteem voor het Vlaamse Gewest)

HSSF Subsurface Flow Wetland with Horizontal Flow

OVAM Public Waste Agency of Flanders (Openbare Afvalstoffenmaatschappij voor

het Vlaamse Gewest)

PE Population Equivalent

SF Surface Flow Constructed Wetland

SS Suspended Solids

SSF Subsurface Flow Constructed Wetland

TN Total Nitrogen

TP Total Phosphorus

TC Total Coliforms

USEPA United States Environmental Protection Agency

VMM Flemish Environment Agency (Vlaamse Milieumaatschappij)

VLAREBO Flemish Soil Sanitation Criteria (Vlaams Reglement Bodemsanering)

VLAREM Flemish Environmental Law (Vlaams Reglement betreffende de

Milieuvergunning)

VSSF Subsurface Flow Wetland with Vertical Flow

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LIST OF TABLES

Table 2.1. General design parameters for SF constructed wetlands (Wood et al., 1995;

De Pauw, 2005)

Table 2.2. General design parameters for SSF constructed wetlands (Wood et al., 1995;

De Pauw, 2005)

Table 2.3. Application of constructed wetlands in developed countries

Table 2.4. Application of constructed wetlands in developing countries

Table 2.5. Oxidation-reduction processes and redox potentials at pH 7 and 25 °C

(Zumdahl, 1992)

Table 2.6. Solubility products of metal sulphides (Zumdahl, 1992)

Table 2.7. Soil remediation criteria (mg kg-1) for nature area and coefficients A, B, C

Table 2.8. Surface water quality criteria (VLAREM II, 2005)

Table 4.2. Detection limits of ICP-OES expressed in µg L-1

Table 5.1. Dissolved Cd, Cr, Ni, and Pb concentration in the wastewater in October 2004

expressed in µg l-1

Table 5.2. Mean removal efficiencies (%) of metals in October and November 2004

Table 5.3. Metal mass removed from the wastewater expressed in kg

Table 5.4. % dry weight in different plant parts

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Table 5.5. Metal concentrations in different plant parts of Phragmites australis in mg kg-1DW

Table 5.6. Concentration of some selected metals in Phragmites australis aboveground plant

parts found in different constructed wetlands expressed in mg kg-1

Table 5.7. Metal mass accumulated in aboveground Phragmites biomass expressed in kg

Table 5.8. Sediment characteristics of the presettlement tank and the surface flow reed bed

Table 5.9. Total metal concentrations in the sediment of the presettlement tank and both

sediment layers of the SF reed bed, with background values and soil remediation

standards, all expressed in mg kg-1 (Vlarebo, 1996)

Table 5.10. Exchangeable metal concentrations in the sediment of the presettlement tank and

both sediment layers of the SF reed bed expressed in mg kg-1

Table 5.11. % of exchangeable metals as a function of sediment depth

Table 5.12. Simultaneously extracted metal concentrations in both sediment layers of the SF

reed bed expressed in mg kg-1

Table 5.13. AVS levels (in µmol g-1) and SEM/AVS ratio in both sediment layers as a

function of distance from the inlet

Table 5.14. Metal mass accumulated in the sediment expressed in kg

Table 6.1. Metal mass removed from the wastewater and metal mass accumulated in the

sediment and Phragmites australis biomass, expressed in kg

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LIST OF FIGURES

Figure 2.1. Comparison between energy inputs of natural and conventional treatment systems

(Kadlec & Knight, 1996; De Pauw, 2004)

Figure 2.2. Different types of aquatic macrophytes (EPA, 2005)

Figure 2.3. Helophyte filter treatment system with surface flow (EPA, 2005)

Figure 2.4. Helophyte filter treatment system with horizontal (HSSF) and vertical (VSSF)

subsurface flow (EPA, 2005)

Figure 4.1. Operational reed bed in Deurle, Flanders (Belgium) – presettlement tank (A) and

a ditch with 9 pipes for receiving discharged wastewater (B) (original, 2004)

Figure 4.2. Schematic presentation of the surface flow constructed wetland in Deurle –

INF = influent, EFF = effluent, Scheidbeek = a creek (original, 2005)

Figure 4.3. Aerial view on the surface flow constructed wetland and sampling positions

(GIS Vlaanderen, 2005)

Figure 4.4. Taking wastewater samples from effluent of the reed bed (original, 2004)

Figure 4.5. Common reed, Phragmites australis (original, 2004)

Figure 4.6. Sediment sampled in polyethylene cylindrical cores (original, 2004)

Figure 4.7. ICP –OES (original, 2004)

Figure 4.8. CEC experiment (original, 2004)

Figure 4.9. Aqua Regia experiment (original, 2004)

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Figure 5.1. Electrical conductivity as a function of sampling position in the constructed

wetland - inf : influent of the presettlement tank; eff PT : effluent of the

presettlement tank; eff: effluent of SF wetland; 0, 25, 70, 140, 350, 630: distance

in meters from the inlet of the SF

Figure 5.2. Dissolved Al (A), Cu (B), and Zn (C) concentration in the wastewater as a

function of sampling position expressed in µg l-1

Figure 5.3. Dissolved Fe (A), and Mn (B) concentration in the wastewater as a function of

sampling position expressed in µg l-1

Figure 5.4. Redox potential of the sediment of the presettlement tank and the upper and

deeper sediment layer in the SF reed bed, expressed in mV

Figure 5.5. AVS of the sediment of the presettlement tank and the upper and deeper

sediment layer in the SF reed bed, expressed in mg S2- kg-1 DM

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LIST OF APPENDICES

Appendix I. Table a – i: Metal concentration in different plant parts as a function of distance

from the inlet expressed in mg kg-1

Appendix II. Figure. a – h: Total metal concentration in both sediment layers of the SF reed

bed as a function of distance from the inlet expressed in mg kg-1

Appendix III. Table a – i: Simultaneously extracted metals (SEM) in both sediment layers of

the SF reed bed as a function of distance from the inlet expressed in mg kg-1

Introduction

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1. INTRODUCTION

One of the environmental problems today is metal pollution of water and soil. The biosphere

is exposed to elevated emissions of metals due to urbanization and industrialization.

Anthropogenic activities may result in the release of metals and represent potential danger to

ecosystems.

Natural wastewater treatment systems include aquatic systems such as facultative ponds and

floating aquatic plant systems, terrestrial systems, and different types of wetland treatment

systems (Kadlec & Knight, 1996; Sundaravadivel & Vigneswaran, 2001). Constructed

wetlands have been used to treat wastewater ever since the pioneering work performed by

Käthe Seidel in the 1960’s (Stottmeister et al., 2003). This technology is suitable for

improving wastewater quality and can be efficiently used for the treatment of various types of

wastewaters. They can provide one of the possible solutions for proper disposal of domestic

wastewaters on a small scale, for example in rural areas or areas where sewage treatment

plants are just too far from households to be connected. Therefore, during the last decade the

number of constructed wetlands in Flanders, the most Northern part of Belgium, increased

exponentially (Rousseau et al., 2004a).

Different types of constructed wetlands are used and they include the following: surface flow

wetlands, subsurface flow wetlands of vertical and horizontal type and their combinations.

The oldest constructed wetland in Flanders is located in Bokrijk. It dates from 1986. It is a

vertical flow reed bed and it is still operational. In Flanders the majority of constructed

wetlands are of the surface flow type and they are mostly planted with common reed

(Phragmites australis (Cav.) Trin. ex steud). Design sizes vary between 1 and 2000

population equivalents (PE), with the majority of reed beds having a size smaller than 500 PE.

Most reed beds are used as single treatment units, although they are sometimes also combined

with other reed beds or even conventional systems. Their main purpose is to treat domestic

and dairy wastewater (Rousseau et al., 2004a).

Different types of wastewater can be contaminated with heavy metals to a different extent.

Concentration levels of trace metals generally do not constitute a major problem in domestic

wastewater, but as they do not degrade, they accumulate in the system and therefore, may

represent a problem after a certain operational lifetime of the wetland (Vymazal et al., 2003;

Introduction

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Lesage et al., 2005). Constructed wetlands are usually open systems, allowing free movement

of biota between the treatment wetland and the adjacent environment. Thus, organisms

exposed to potentially dangerous levels of metals in wetland treatment systems may move

offsite and contribute to the contamination of natural areas or become part of the human food

chain (Kadlec & Knight, 1996).

This thesis aimed at assessing heavy metal accumulation in a surface flow wetland.

Investigated surface flow wetland is one of the oldest constructed wetlands in Flanders, and

has been in operation for 16 years. Although the surface flow wetland treats domestic

wastewater in which metal concentration levels are low, the question arose whether after this

period of time elevated metal concentrations occurred in the sediment and plants and whether

there was a potential risk. The distribution of heavy metals over different biotic and abiotic

wetland compartments (wastewater, Phragmites australis aboveground plant parts, and

sediment) was therefore investigated. Heavy metal concentrations in the influent and effluent

wastewater were analysed in order to assess the removal efficiency of the constructed

wetland. Total metal concentrations in the sediment were determined to get an idea of the

pollution level. However, total concentrations in the sediment are not a good indicator of the

availability of the metals and are not useful to determine potential risks (Meers, 2005). An

idea of the potential mobility of metals in the sediment of the surface flow wetland was

estimated by means of the SEM/AVS ratio and by means of the NH4OAc-extractable metal

concentrations, representing the exchangeable metal fraction in the sediment.

Literature review

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2. LITERATURE REVIEW

Different types of natural wastewater treatment systems will be discussed in the following

section. Special attention is paid to surface flow treatment wetlands and an overview of their

advantages and disadvantages is presented. As well, removal processes of heavy metals and

their accumulation in different biotic and abiotic compartments of surface flow wetlands will

be discussed.

2.1. Natural wastewater treatment systems as an alternative to conventional

wastewater treatment systems

Conventional wastewater treatment relies on large-scale plants and it is the preferred form of

wastewater treatment in developed countries. Minimizing the area required for treatment per

capita, it is an important consideration in urban areas where land space is limited. An

additional advantage of conventional treatment systems is that they can treat more wastewater

over a certain period of time because the retention time of the wastewater is shorter. However,

an important drawback of conventional systems is that they require energy and are therefore

costly compared to natural systems. Natural systems may be used as an interesting alternative.

Man today tends to turn to natural systems for wastewater treatment especially in small rural

areas where the cost would be too high to connect to the central sewer system (Rousseau et

al., 2004a, Sundaravadivel & Vigneswaran, 2001). Prerequisites for natural treatment

systems, called low rate systems as well, are the presence of sufficient light, not too low

temperature, and a wastewater that is not toxic. The detention time must be long enough,

varying from days to weeks or months, which is quite long compared to conventional systems

where it is only a matter of hours. The organic loading, determined by volume and strength of

the wastewater, should not be too high and pretreatment is needed in order to sufficiently

reduce BOD (De Pauw, 2004).

The fossil fuel intensiveness of conventional treatment is a major disadvantage compared to

sun and wind driven natural systems. Natural systems are considered a "green" and

sustainable technology because they require less non-renewable energy sources than other

alternatives (Brix, 1999). Their major disadvantage is their land intensiveness (Fig. 2.1.).

Literature review

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Figure 2.2. Comparison between energy inputs of natural and conventional treatment systems

(Kadlec & Knight, 1996; De Pauw, 2004)

Natural processes have always cleaned water when it flows through rivers, lakes, streams, and

wetlands. In the last several decades, systems have been constructed to use these natural

processes for water quality improvement and to reduce different kinds of pollutants.

Plant systems are very useful to humans in keeping sustainable development of a certain area.

According to Kadlec & Knight (1996) natural wastewater treatment systems can be

categorized into three major categories: (1) aquatic or pond/lagoon systems, (2) terrestrial or

land application systems, and (3) wetland systems. These treatment systems will be discussed

in the following paragraphs. Aquatic and terrestrial systems will be discussed briefly, whereas

special attention will be paid to constructed wetlands.

2.1.1. Aquatic wastewater treatment systems

Natural aquatic treatment systems involve impounding wastewater in ponds or lagoons for a

sufficient period so that pollutants and pathogens in wastewater are removed through natural

biological degradation processes. Wastewater stabilization ponds (WSPs) are simple, low

maintenance and low cost systems that treat wastewater (De Pauw, 2004). Different types of

aquatic systems exist.

Literature review

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Shallow man-made basins consisting of a series of anaerobic, facultative or maturation ponds

represent waste stabilization ponds (Nelson et al., 2003). In the anaerobic pond the primary

treatment takes place, mainly designed for removing suspended solids and BOD. Most of the

remaining BOD is removed in the facultative pond through the combined activity of algae and

heterotrophic bacteria. The main function of the tertiary treatment in the maturation pond is

the removal of pathogens and nutrients (especially nitrogen). These systems are suitable for

tropical and subtropical countries because the intensity of the sunlight and temperature that

are the key factors for the efficiency of the removal processes.

Stabilization ponds can be aerated by mechanical systems such as surface-type aerators or

submerged propeller-type aerators, designed to provide sufficient oxygen for biological

degradation processes and an adequate mixing. Nelson et al. (2003) disccussed about four

primary wastewater stabilization ponds in central Mexico (three facultative and one

anaerobic). Many other authors mention usage of stabilization ponds (Garcia et al., 2000;

Sinton et al., 2001; Oakley et al., 2000; etc…).

2.1.2. Terrestrial wastewater treatment systems

Terrestrial wastewater treatment systems use an unsaturated soil layer to provide both direct

filtration and assimilation of pollutants and a rooting medium for plant growth that aids in the

filtration and uptake of pollutants from wastewater (Sunduravadivel & Vigneswaran, 2001).

Houlbrooke et al. (2004) discusses a nutrient-rich farm-dairy effluent (FDE) treatment system

in New Zealand. Terrestrial treatment systems include on site infiltration systems, slow rate

land application systems, high rate land application systems, and overland flow systems

(Kadlec & Knight, 1996). The main difference is that on site infiltration and land application

systems discharge the treated effluent into the groundwater (‘zero discharge systems’), while

overland flow systems discharge to surface waters.

On site infiltration systems include, for example, residential septic tanks. Slow rate land

application systems are using irrigation of vegetated systems for wastewater disposal (Kadlec

& Knight, 1996). They are characterized by low loading rates, low oxygenation rates, and

high volume and area requirements (De Pauw, 2004). High rate land application systems are

using highly permeable soil for groundwater discharge. Overland flow systems rely on sloped

Literature review

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vegetated terraces with impermeable soils to restrict infiltration with wastewater and to direct

the wastewater down to collection channels (Kadlec & Knight, 1996).

2.1.3. Constructed wetlands

Wetlands are sometimes described as ‘kidneys of the landscape’ due to their ability to

transform and store organic material and nutrients (Brix, 1994). Although wetlands have been

created incidentally by human and animal engineering of ponds and lakes throughout the

history, the intentional construction of wetlands to provide habitat and/or water quality

functions began with the environmental movement in the 1970’s (Kadlec & Knight, 1996).

According to the United States Environmental Protection Agency (EPA, 2005) constructed

wetlands can be defined as:

‘engineered or constructed wetlands that utilize natural processes involving wetland

vegetation, soils and their associated microbial assemblages to assist, at least partially, in

treating an effluent or other waste source’.

This cheap and effective wastewater treatment technology may be used in temperate and

tropical climates, in developed and developing countries as will be discussed in paragraph

2.1.3.4. Wetlands may have several functions such as: habitat wetlands, water treatment

wetlands, flood control wetlands, and aquaculture wetlands. One may state that constructed

wetlands are wetlands intentionally created from non wetland sites for the sole purpose of

wastewater or storm water treatment, whereas created wetlands are intentionally created from

non wetland sites to produce or replace natural habitat. Flood control wetlands are

impoundments used to offset losses of natural flood storage volumes by urban and

agricultural development, while constructed aquaculture wetlands are systems integrating

aquaculture or crop production with wastewater treatment (Kadlec & Knight, 1996).

Literature review

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2.1.3.1. Classification of Constructed Wetlands

Constructed wetlands can be classified according to the life form of the dominating

macrophyte in the wetland (Fig. 2.2.) (Brix, 1994, De Pauw, 2003):

• Emergent macrophyte systems also named helophyte filters: Phragmites sp., Typha

sp., Scirpus sp., Zizania aquatica

• Free floating macrophyte systems also named pleustophyte filters: Nymphaea sp,

Nuphar spp., Eichornia crassipes, Lemna spp., Azolla sp.

• Submerged macrophyte systems also named hydrophyte filters: Elodea sp.,

Ceratophyllum sp., Isoetes sp.

Figure 2.2. Different types of aquatic macrophytes (EPA, 2005)

Many different types of constructed wetlands are used in Flanders, ranging from surface over

subsurface of horizontal and vertical type and other types with all possible combinations. A

review of different types of helophyte filters will be presented in the following section.

Rousseau et al. (2004a) mentioned 107 constructed wetland systems since 1986, of which 54

are free surface flow wetlands. The design of SF constructed wetlands in Flanders varies

between 1 PE and 2000 PE with the majority having a capacity smaller than 500 PE, and an

average footprint area of about 7 m² PE-1. An average investment cost of € 392 PE-1 is

described. Average removal efficiencies of SF reed beds are the lowest ones (COD 61 %; SS

75 %; TN 31 % and TP 26 %) mainly due to the diluted influent from the combined sewer

systems and the limited contact with the soil or filter medium (Rousseau et al., 2004a).

Literature review

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2.1.3.2. Types of helophyte filters

Surface flow constructed wetlands (SF)

Flooded systems where wastewater is exposed to the atmosphere are called surface flow

wetlands (Fig. 2.3.). Surface flow wetland cells are mostly designed as rectangular-shaped

basins with the inlet and outlet located on opposite sides of the system. Design of constructed

wetlands includes building of multiple cells and each cell should provide the same level of

treatment. The size and number of elements of surface flow systems are based on estimates of

strengths of the influent wastewater. Hydrology and pollutant removal processes that occur in

the wetland are important for the design and operation of a constructed wetland. The water

balance and detention time are important as well. Surface flow wetlands are subject to losses

(evapotranspiration and seepage) and gains (rainfall), which lead to fluctuating volumes and

levels within the wetland. A sufficiently long detention time is important for adequate

treatment and varies between 2-5 days for BOD and 7-14 days for nitrogen removal (Crites et

al., 1997). Surface flow systems can significantly reduce biological oxygen demand (BOD),

chemical oxygen demand (COD) and suspended solids (SS) (EPA, 1988). Their treatment

capacity is about 1000 PE ha-1.

Figure 2.3. Helophyte filter treatment system with surface flow (EPA, 2005)

Precipitation, infiltration, evapotranspiration, hydraulic loading rate, and water depth can all

affect the removal of organic pollutants, nutrients, and trace metals, not only by altering the

detention time, but also by concentrating or diluting the wastewater (EPA Manual, 1988). The

Literature review

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shallow water depth, low flow velocity and presence of plants regulate water flow and,

especially in long, narrow channels, ensure plug-flow conditions (EPA Manual, 1988). Cells

are first excavated and the bottom is designed in that way that it has a slight downgrade slope

(approximately 0.5 percent), assisting the flow of wastewater through the cells by gravity. The

bottom of each cell should be provided with a liner such as clay, bentonite or a synthetic liner.

This prevents leakage and possible contamination of groundwater and the surrounding

environment. On top of the liner soil is placed to support plant growth (Hoban et al., 2005).

To prevent short-circuiting of the wastewater flow, and to make the best use of every cm of

cell space, the wastewater in constructed wetlands should be evenly distributed across the

width of each cell (Brix, 1994). Perforated distribution pipes serve as an entrance for

wastewater to surface flow wetlands.

General design parameters are presented in Table 2.1. and include the detention time, the

aspect ratio, etc… The ratio of the cell length to its width, the aspect ratio, usually ranges

from 2:1 to 4:1, but may be higher depending on the site and other factors (Hoban et al.,

2005). Other authors state that the length to width (L/W) ratio is a key factor to achieve

adequate wastewater treatment and that it should be at least 10 (Sundaravadivel &

Vigneswaran, 2001). Depth of water column is about 0.4 m.

Table 2.1. General design parameters for SF constructed wetlands

(Wood et al., 1995; De Pauw, 2005)

Factor Typical SF Units

Detention time 5 - 14 days

Max BOD loading rate 80 kg/ha/day

Water or substrate depth 0.1 - 0.5 m

Hydraulic loading rate 7-60 mm/d

Areal requirement 0.002 - 0.014 ha/m³/day

Aspect ratio (L/W) 2/1 – 10/1 -

Mosquito control Required -

Harvest frequency 3 - 5 year

The placement of the plants can be planned and arranged. For example, some surface flow

cells are designed to have areas of open water as well as areas of dense vegetation to allow

wind and sunlight to reach parts of the cell to influence flow and treatment.

Literature review

10

Most systems are designed for the wastewater to flow once through the system. However,

systems can be designed to recirculate all or a portion of the effluent to treat the wastewater

more than once and improve the effluent quality (Brix, 1994). Surface flow wetlands have

few maintenance requirements, but maintenance must be performed properly to ensure system

performance. Examples of maintenance are alternating cells or adjusting water levels and

harvesting the vegetation (Kadlec & Knight, 1996). Some systems may have banks that need

to be maintained, and inlet and outlet structures that should be cleaned periodically.

Subsurface flow constructed wetlands (SSF)

Subsurface flow systems have the water level below the surface of the medium placed in the

beds. Systems with subsurface horizontal flow maintain the medium water saturated, whereas

in vertical flow systems the medium is not saturated, because the water is usually applied at

timed intervals and allowed to percolate through the medium (Fig. 2.4.).

Figure 2.4. Helophyte filter treatment system with horizontal (HSSF) and vertical (VSSF)

subsurface flow (EPA, 2005)

Subsurface flow wetlands with horizontal flow

In a SSF wetland with horizontal flow, the basin is usually filled with gravel or another coarse

substrate. General design parameters are presented in Table 2.2. The depth of the bed is

usually less than 0.6 m and the bed bottom is sloped in order to minimize flow above the

Literature review

11

surface. The treatment capacity varies between 2000 – 4000 PE ha-1 (Sundaravadivel &

Vigneswaran, 2001; Vymazal, 2002).

Table 2.2. General design parameters for HSSF constructed wetlands

(Wood et al., 1995; De Pauw, 2005)

Factor Typical SSF Units

Detention time 2 - 7 days

Max BOD loading rate 75 kg/ha/day

Water or substrate depth 0.10 – 1 m

Hydraulic loading rate 2 - 30 mm/d

Areal requirement 0.001 - 0.007 ha/m³/day

Aspect ratio (L/W) 0.25/1 – 5/1 -

Mosquito control Not required -

Harvest frequency 3 - 5 year

Subsurface flow wetlands with vertical flow

These systems are also called ‘infiltration wetlands’ due to the infiltration of wastewater

through the substrate bed. The depth of the bed is 2-3 m. Water purification happens mainly

through the substrate and associated microorganisms, microorganisms that colonize plant

roots, and the plant itself. The plants provide better percolation of water through the filter

substrate. Plants raise nitrification by giving oxygen in the rhizosphere (De Pauw & De

Maeseneer, 1992; Colmer, 2003). The treatment capacity may be 5000 PE ha-1 (De Pauw &

De Maeseneer, 1992).

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12

2.1.3.3. Advantages & disadvantages of constructed wetlands

Each technology has weak and strong sides that have to be considered before and while

planning and implementing a system. The following paragraph presents some of the

advantages and disadvantages of constructed treatment wetlands.

Advantages of constructed wetlands represent improving water quality and providing

effective secondary or tertiary treatment, removing different kinds of pollutants and

pathogens. They can be used for the treatment of different types of wastewater, such as

domestic wastewater (Rousseau et al., 2004a), acid mine drainage (AMD) (Mitsch et al.,

1997; Mays et al., 2000; Hallberg et al., 2005), urban runoff (Scholes et al., 1998), and storm

water (Bavor et al., 2001). Solano et al. (2004) described the high removal efficiency of total

coliforms (TC), faecal coliforms (FC) and faecal streptococci (FS) in reed beds treating

domestic wastewater in rural areas.

Constructed wetlands are inexpensive systems, having low investment and maintenance costs

and requiring little or no energy to operate. They can be operated by relatively untrained

personnel rather than by high expert knowledge (Solano et al., 2004). Constructed wetlands

can provide additional wildlife habitat and can be aesthetically pleasing environments to

neighborhoods. They can also have an educational value.

Kivaisi (2001) mentions that most of the developing countries have warm tropical climates

that are beneficial for a high biological activity and productivity, making constructed wetlands

an interesting treating option in developing countries. The same author points at the secondary

function of the biomass as a fuel.

On the other side, disadvantages of constructed wetlands are that they require more land area

than many other treatment options. However, this usage of land is less intensive than in other

systems. An additional disadvantage may be that wetlands are not appropriate for treating

wastewater with high concentrations of certain pollutants. The performance of wetlands may

also vary based on usage and climatic conditions and there may be a long initial start-up

period before the vegetation is adequately established.

Literature review

13

Finally constructed wetlands can often be described as ‘mosquito friendly habitats’ and may

raise potential conflicts with neighboring human population (Knight, 2003). Mosquitoes are

possible vectors for viral infections (such as malaria, filariasis, encephalitis) so special care

must be given to integrate mosquito control into the design of wetland in tropical areas

(Kivaisi, 2001). According to Johansson et al. (2004) wetlands are methane sources and may

contribute to increasing levels of this greenhouse gas. Methane has a global warming potential

(GWP) 21 times higher than CO2 (Tchobanoglous et al., 1994).

2.1.3.4. Usage of constructed wetlands in developed and developing countries

Due to a better financial situation and more stringent environmental standards, environmental

laws are much more developed in industrialized countries. Conventional wastewater treatment

systems are widely used in developed countries but natural wastewater treatment systems may

represent efficient additional treatment systems or alternatives. Vymazal et al. (2001)

mentions the use of constructed wetlands for wastewater treatment in European countries such

as: Austria, Belgium, Czech Republic, Denmark, France, Germany, Hungary, Norway,

Poland, Portugal, Slovenia, Sweden, Switzerland, the Netherlands and the United Kingdom.

In Australia and the US (Kadlec & Knight, 1996) these systems are in use as well. Table 2.3.

presents some developed countries where constructed wetlands are currently used with their

references.

Table 2.3. Application of constructed wetlands in developed countries

Country Reference

Austria Haber et al. (1998)

Australia Bavor et al. (2001)

Belgium Rousseau et al. (2004)

Germany Rustige et al. (2001); Platzer et al. (1994)

Denmark Schierup et al. (1997)

Ireland Otte (2004)

The Nederlands Verhoeven et al. (1998)

New Zealand Tanner et al. (2003)

UK Scholes (1998)

USA Kadlec & Knight (1996); Mays et al. (2000)

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14

As the population keeps growing, especially in developing countries, there is a growing need

for cost effective wastewater treatment. Liquid wastes (untreated sewage or industrial waste)

are a major source of pollutants in developing countries and quite often they are discharged

into aquatic environments without any treatment (Kivaisi, 2001). Costs of building

conventional wastewater treatment plants are simply too high for developing countries.

Natural treatment systems could be introduced as an environmentally friendly solution. As an

inexpensive, low-maintenance technology, it is becoming increasingly in demand in countries

of Central and South America, Eastern Europe and Asia. Table 2.4. presents some developing

countries where constructed wetlands are currently used with their references.

Table 2.4. Application of constructed wetlands in developing countries

Country Reference

China Ye et al. (2001)

Costa Rica Dallas et al. (2004)

Czech republic Vymazal (2001); Žakova (1996)

India Juwarkar et al. (1995); Billore et al. (1999)

Kenya Nyakang’o (1999)

Nepal Shreshta et al. (2000)

Poland Kufel (1991)

Tanzania Kivaisi (2001)

Literature review

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2.2. Heavy metal accumulation in constructed wetlands

The main heavy metals of concern in water quality management are lead, copper, zinc,

chromium, cadmium, mercury, and arsenic. These metals may be highly toxic if present in

elevated concentrations in water (Sundaravadivel & Vigneswaran, 2001). Metals are

persistent and can accumulate in constructed wetlands even when low concentrations are

present in the wastewater. Metals are mostly accumulated in the sediment of constructed

wetlands (Vymazal, 2003; Lesage et al., 2005). The total metal concentration in the sediment

gives information on the pollution level, but does not give a good indication of the mobility

and ecotoxicity of the metals. Metals exist in different physicochemical forms, affecting their

mobility and availability. Metals in sediments or soils can exist in the following forms

(Meers, 2005):

• free metal ions and soluble metal compounds in the soil solution

• exchangeable metal ions sorbed onto inorganic solid phase surfaces

• non-exchangeable metal ions, either present as precipitates or insoluble inorganic

metal compounds (eg. oxides, hydroxides, phosphates, carbonates, sulphides)

• metals complexed by soluble or insoluble organic materials

• metals incorporated in the clay crystalline structure.

Different forms (or species) of metals have different availability, different rates of uptake and

different effects. Water soluble and exchangeable metals are the most available. Potentially

available metals are metals precipitated as inorganic compounds, metals complexed with large

molecular weight humic materials and metals adsorbed to hydrous oxides. Unavailable forms

are metals bound within the crystalline lattice of minerals (Weiss, 2004).

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2.2.1. Removal processes of heavy metals in surface flow constructed wetlands

Pollutants are removed through a combination of physical, chemical and biological processes

including sedimentation, precipitation, adsorption to sediment and soil particles, assimilation

by plant tissues and microorganisms (Watson et al., 1989; Brix, 1994).

There are three main removal processes that remove heavy metals in surface flow wetlands:

• binding processes to soils, sediments, and particulate matter

• precipitation as insoluble salts

• uptake by microorganisms, algae, and plants.

Due to their positive charge, metals are readily adsorbed, complexed and bound with

suspended particles, which are removed via sedimentation and filtration (Sundaravadivel &

Vigneswaran, 2001).

When precipitation occurs, insoluble salts are created such as sulphides, hydroxides,

carbonates and bicarbonates and metals are fixed into the wetland sediment (Sundaravadivel

& Vigneswaran, 2001).

It is known that algae and microorganisms take up heavy metals available in the dissolved

form, whereas macrophytes can also take up metals from the sediments. During the initial

period of establishment of wetlands, the binding processes are limited and the uptake by the

biota is dominant (Sundaravadivel & Vigneswaran, 2001).

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17

2.2.2. Factors influencing metal mobility

Sediments play an important role in element cycling in the aquatic environment. They

mediate their uptake, storage, release and transfer between environmental compartments.

Important abiotic factors that influence metal mobility in surface flow wetland sediments are

the pH, the oxidation-reduction status (redox potential or Eh), the amount of organic matter,

and the texture of the sediment. Biological factors can affect metal mobility by influencing

above mentioned factors. A review of influencing factors is presented.

Influence of texture

The texture is determined by the particle size distribution. A differentiation between 3

fractions is made: sand (> 50 µm), silt (2 – 50 µm), and clay (< 2 µm). Sand is composed of

large, neutral particles of silicon dioxide (SiO2). Sandy soils usually have a low cation

exchange capacity (CEC) and are therefore less capable of retaining metals than clay soils.

Chemically, clays are aluminosilicates [Al4Si4O10(OH)8] and carry negative charges

(Gambrell, 1994). Clay is an essential component of a productive soil. It plays a vital role in

holding plant nutrients and water. Because clays have a large surface area and negative

charges, they can attract and hold positively charged ions. This characteristic is important

because many positively charged ions are plant nutrients, such as calcium, magnesium, and

potassium. Verloo (2004) showed that soil texture plays an important role in the mobility of

Zn. It showed that the sandier the soil, the better the solubility at higher pH.

Influence of redox potential

The oxidation-reduction status of a sediment is expressed as a potential (the redox potential

Eh, in mV) and is a measure for the electron availability. Aerobic soils have high redox

potential whereas water saturated soils and sediments usually have low redox potentials.

Water saturated soils have limited gaseous oxygen diffusion and available dissolved oxygen is

consumed as a terminal electron acceptors by microbial respiration. When dissolved oxygen

becomes limiting, a sequence of reduction processes takes place with a decrease of the redox

potential as a result (Gambrell, 1994). When dissolved oxygen is consumed, microorganisms

turn to nitrate, oxidized forms of Fe and Mn, sulphates, and finally CO2 as an electron

acceptor (Table 2.5.).

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Table 2.5. Oxidation-reduction processes and redox potentials at pH 7 and 25 °C (Zumdahl,

1992)

Reaction Eh (mV)

Reduction of O2

O2 + 4 H+ + 4 e- ↔ 2 H2O

812

Reduction of NO3-

2 NO3- + 12 H+ + 12 e- ↔ N2 + 6 H2O

747

Reduction of Mn4+

Mn4+ + 4 H+ + 2 e- ↔ Mn2+ + 2 H2O

526

Reduction of Fe3+

Fe3+ + 4 H+ + 8 e- ↔ Fe2+ + 2 H2O

-47

Reduction of SO42-

SO42- + 10 H+ + 8 e- ↔ H2S + 4 H2O

-221

Reduction of CO2

CO2 + 8 H+ + 8 e- ↔ CH4 + 2 H2O

-244

When there is no dissolved oxygen and nitrate present, microorganisms turn to Mn4+ and Fe3+

as terminal electron acceptors as mentioned, reducing them to ferrous iron (Fe2+) and

manganous manganese (Mn2+). Those are more soluble and available to plants. As a result of

reducing conditions, Fe and Mn were found to accumulate in plants (Gambrell, 1994).

Reduction of Fe- and Mn-oxides can lead to the release of co-precipitated heavy metals and

an increase of their mobility.

When the redox potential continues to decrease, as is the case in sediments that are water

saturated during prolonged periods of time, sulphides may be formed and precipitate most of

the heavy metals as CuS, PbS, CdS, etc… (Verloo, 2004). The solubility product of these

metal sulphides is very low, leading to a low availability of metals in these forms.

Influence of pH

The sediment pH is a major factor influencing metal chemistry and mobility. Natural wetland

soils have a pH range of 6.5 – 7.5 and the near neutral pH conditions favor metal

immobilization (Gambrell, 1994). Generally metal solubility and mobility increase when the

pH decreases and the sediment gets more acid. This can happen when reduced sediments

become oxidized.

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19

Influence of organic matter

Large inputs of decomposing organic matter are present in surface flow wetlands, such as

organic matter from the influent wastewater and decaying Phragmites australis biomass.

Decomposing organic matter forms humus components, polyelectrolytes that form complexes

with metals. Humus components are grouped into humic acids (soluble in alkali environment)

and fulvic acids (soluble in acid and alkali environment) (Verloo, 2004). Fulvic acids can

form soluble complexes with heavy metals, thereby increasing their mobility. Complexes

formed with humic acids are generally insoluble especially in acid media and can therefore

represent a sink for heavy metals (Verloo, 2004). Increasing the organic matter content of any

soil will help to increase the CEC since it also holds cations.

Influence of Phragmites australis

Plants do not only affect the soluble metal concentration in the sediment by the direct uptake

of metals by the roots, plants can also directly affect certain sediment characteristics and

thereby influence the metal mobility in the sediment.

Phragmites australis plants have the ability to oxidize the sediment in the root zone through

the release of O2 by the aerenchyma tissue. If this happens in strongly reduced sediment, this

oxygen release by plant roots can lead to the oxidation of sulphides and the mobilization of

metals (Weiss, 2004).

Plant roots can excrete organic acids that increase the metal mobility by increasing the pH in

the rhizosphere and forming complexes with metals (Poschenrieder & Barcelo, 2002).

Organic acids serve a role as a natural chelating agent.

When plants are not harvested at the end of the growing season, large amount of organic

matter are introduced into the wetland system (Gessner, 2000). This decomposing organic

matter enlarges the CEC of the sediment and thereby affects the metal mobility.

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20

2.2.3. Estimating metal mobility

Single extraction procedures

The total amount of metals does not give information on their mobility. Metals can occur in

different physicochemical forms, influencing their mobility and availability. Sequential and

single extraction procedures are used to classify the chemical forms into operationally defined

fractions. Extraction of the sediment with a certain reagent allows to estimate the distribution

over the different fractions and to make an estimation of the mobility and availability. Water

extractable metals represent the metals in the soil solution. The exchangeable fraction consists

of metals that can be readily released into the soil solution by cation exchange processes.

Proposed extractants to assess this fraction are CaCl2, MgCl2, NH4NO3, NaNO3, KNO3,

Mg(NO3)2, NH4OAc (Meers, 2005). Water extractable and exchangeable metals represent

labile metal fractions and give an idea of the metal mobility in the sediment.

SEM/AVS ratio

Metal mobility in reduced sediments of surface flow wetlands, is strongly affected by the

presence of sulphides. Acid volatile sulphide (AVS) is a measure for the amount of metal

sulphides in the sediment and is defined by the amount of sulphide released after extracting

the sediment with 1 M HCl. Metals precipitate with sulphides according to the solubility

products (Ksp) presented in the Table 2.6. The metal sulphide with the lowest solubility

product will be the most stable precipitate.

Table 2.6. Solubility products of metal sulphides (Zumdahl, 1992)

Metal sulphide Ksp

MnS 2.3 × 10-13

FeS 3.7 × 10-19

ZnS 3.0 × 10-21

NiS 2.5 × 10-22

CdS 1.0 × 10-28

PbS 7.0 × 10-29

CuS 8.5 × 10-45

Literature review

21

Metals precipitated with sulphides have a very low availability and therefore form a small risk

when it comes to toxicity. SEM are the metals released together with the extraction of the

AVS. SEM is a parameter that is however operationally defined and can also contain metals

present in the other forms (such as carbonates, oxides, organic matter). Thus, the extraction

with 1 M HCl will not only release metals from the sulphides, but can also extract metals

from other fractions as well. The SEM/AVS ratio can be considered as a measure for the

potentially bioavailable metal fraction (van den Hoop et al., 1997). When the SEM/AVS ratio

is larger than 1 this means that metals are present that are not bound by sulphides and are thus

potentially bioavailable. However, as the extraction with 1 HCl is not selective, this does not

mean that those metals are mobile, they can still be bound by other components such as

organic matter and carbonates. It gives an indication, and can therefore be described as a

measure of potential metal mobility.

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2.3. Belgian Legislation Framework

2.3.1. Soil remediation criteria (Vlarebo, 1996)

Background levels are total concentrations in soils, which are not affected by human

activities. Background levels and soil remediation criteria are described for a standard soil

with 2% organic matter and 10% clay. Site-specific criteria are postulated, based on the

organic matter and clay content according to the following formula:

A + B * x + C * y

N (x, y) = N (10, 2) * ------------------------

A + B * 10 +C * 2

With

N: background value or soil remediation criterion with clay content x % or 10 %

and organic matter content of y % or 2 % (in mg kg-1 DW)

A, B, C: coefficients, depending on type of metal

x: clay content of the soil (%)

y: organic matter content of the soil (%)

Background values for a standard soil are set at 19 mg kg-1 for As, 0.8 mg kg-1 for Cd, 37 mg

kg-1 for Cr, 17 mg kg-1 for Cu, 0.55 mg kg-1 for Hg, 40 mg kg-1 for Pb, 9 mg kg-1 for Ni, and

62 mg kg-1 for Zn. Coefficients A, B, and C are presented in Table 2.7. Soil remediation

criteria depend on the land use (nature, agricultural, residential, recreational, and industry). In

Table 2.7. the most stringent soil remediation criteria (for a nature area) for a standard soil

with coefficients A, B, and C are presented as well.

Literature review

23

Table 2.7. Soil remediation criteria (mg kg-1) for nature area and coefficients A, B, C

Element N(10,2) A B C

As 45 14 0.5 0

Cd 2 0.4 0.03 0.05

Cr 130 31 0.6 0

Cu 200 14 0.3 0

Hg 10 0.5 0.0046 0

Pb 200 33 0.3 2.3

Ni 100 6.5 0.2 0.3

Zn 600 46 1.1 2.3

2.3.2. Surface water quality criteria

As of 1 July 1995, the quality criteria presented in Table 2.8. apply for all surface waters in

Flanders. 90 % of the measurements in one year must comply with this standard. Of the 10 %

that does not comply, the wastewater may not have a deviation larger than 50 % from the

standard.

Table 2.8. Surface water quality criteria (VLAREM II, 2005)

Element Criteria (µg L-1)

Cu (tot) 50

Pb (tot) 50

Zn (tot) 200

Cr (tot) 50

Ni (tot) 50

As (tot) 30

Fe (dis) 200

Mn (dis) 200

Se (tot) 10

Ba (tot) 1000

Cd (to) 1

Hg (tot) 0.5

Objective of the study

24

3. OBJECTIVE OF THE STUDY

The objective of this study is to assess metal accumulation in a free surface flow wetland used

for the treatment of domestic wastewater. The constructed wetland is located in Deurle,

Flanders and has been in operation since 1989. The wetland is one of the oldest constructed

treatment wetlands in Belgium. As heavy metals are persistent, they accumulate in the

constructed wetland and can have implications on the operational lifetime of the system and

present a possible threat to the ecosystem. Goal of the presented study was therefore to

perform a preliminary investigation towards heavy metal accumulation in different biotic and

abiotic compartments of this relatively old surface flow constructed wetland and to assess

potential risks. Three compartments of the constructed wetland (water, sediment and

aboveground plant parts of Phragmites australis plants) were sampled and analysed for heavy

metal concentrations.

The removal of heavy metals in the surface flow wetland was investigated by sampling

influent and effluent wastewater. Water samples were also taken at intermediate sampling

positions in order to study the removal pattern of heavy metals along the length of the

constructed wetland. Phragmites australis aboveground plant parts were sampled at the same

intermediate sampling positions in order to assess heavy metal concentrations in leaves,

stems, leaf sheaths and panicles. Sediment was sampled at intermediate sampling positions in

the surface flow wetland and at the inlet and outlet area of the presettlement tank. General

physicochemical sediment characteristics were determined. The pollution level of the

sediment was assessed by analysing total metal concentrations. However, total metal

concentrations in the sediment are not a good indicator of the mobility and availability of

metals and are not useful to determine potential risks. The potential mobility of metals in the

sediment of the surface flow wetland was estimated by means of the SEM/AVS ratio and by

means of the NH4OAc-extractable metal concentrations, representing the exchangeable metal

fraction in the sediment. The possible effect of depth was investigated by studying two

sediment layers, 0-15 cm and 15-30 cm deep.

Materials and methods

25

4. MATERIALS AND METHODS

4.1. Study site

The constructed surface flow wetland is situated in the village of Deurle, part of Sint Martens-

Latem, 15 km southwest from Ghent. It is designed to treat domestic wastewater of the village

of Deurle, with a design capacity of 900 PE. The constructed wetland was designed and

constructed by the VLM (Vlaamse Landmaatschappij - the Flemish Land Agency), but was

handed over to Aquafin, the state owned company for wastewater treatment. The constructed

wetland has been in operation since 1989. It is situated in a nature area (Gisvlaanderen, 2005).

The reed bed is shown in Figure 4.1. and a schematic presentation of the constructed wetland

is presented in Figure 4.2.

A B

Figure 4.1. Operational reed bed in Deurle, Flanders (Belgium) – presettlement tank (A) and

a ditch with 9 pipes for receiving discharged wastewater (B) (original, 2004)

The domestic wastewater is introduced via two pumps into the presettlement tank and then

flows through the surface flow reed bed after which it is collected in a discharge ditch. The

discharge ditch has a length of 48 m and a width of 2 m. Then effluent wastewater flows

gravitationally into the Scheidbeek, a creek nearby (Leuridan, 2004). Creek Scheidbeek later

on enters the tourist-attractive river Lys – Leie (Creele, 1992).


26

Figure 4.2. Schematic presentation of the surface flow constructed wetland in Deurle –

INF = influent, EFF = effluent, Scheidbeek = a creek (original, 2005)

The presettlement tank is 51 m long and has a trapezoidal cross section with a bottom width

of 2 m and a width of 4 m at the water surface. The surface flow reed bed has a length of 85

and width of 43m. There is no liner present. The reed bed consists of 9 channels and it is

planted with Phragmites australis. Each channel has a trapezoidal cross section with a bottom

width of 2 m and a width of 4 m at the water surface. A water depth of about 0.5 m is

observed in the channels.

The way the water has flown through the channels in the past is not registered and appears to

have occurred in an uncontrolled manner (Kluft, 2005). At the time of sampling the

wastewater flowed through the channels in series.


27

4.2. Sampling

In September 2004, wastewater, sediment, and vegetation samples were taken from the

surface flow reed bed. Wastewater was sampled a second time in October 2004. Six

intermediate sampling locations were chosen along the length of the constructed wetland: at 0,

25, 70, 140, 350, and 630 m from the inlet (Fig. 4.3). Wastewater, sediment, and aboveground

plant parts were sampled at these intermediate locations. Sediment and wastewater were also

sampled at the inlet and outlet area of the presettlement tank. A sample of the effluent

wastewater was also collected.

Figure 4.3. Aerial view on the surface flow constructed wetland and sampling positions

(GIS Vlaanderen, 2005)

4.2.1. Wastewater

Polyethylene bottles were used for sampling of wastewater. Samples were taken in three

replicates at each intermediate sampling position, from the influent of the presettlement tank,

and from the influent and effluent of the reed bed. During the first sampling in September

2004, samples were taken by hand by wading into the surface flow wetland, whereas during

the second sampling in October 2004 samples were taken by throwing a polyethylene bottle

Surface flow reed bed

Presettlement area

Sampling point

Effluent collection

Surface water


28

attached to a rope into the water from the surrounding banks. Samples were stored in

polyethylene beakers and conserved at 4°C until analysis (Fig. 4.4.).

As the sediment surface was disturbed during the first sampling, samples were filtered over

0.45 µm Millipore filter paper. In each sample three drops of 65 % HNO3 was added in order

to conserve the samples. Dissolved heavy metal concentrations, pH, and electrical

conductivity were determined in the wastewater samples.

Figure 4.4. Taking wastewater samples from effluent of the reed bed (original, 2004)

4.2.2. Aboveground Phragmites australis biomass

Aboveground plant parts of Phragmites australis were cut with scissors above the water

surface at the six intermediate sampling positions in the constructed wetland. Three replicate

plants were sampled at each position. Plants were conserved in polyethylene bags during

transportation to the laboratory (Fig. 4.5.).

In the laboratory plants were differentiated into stems, leaves, leaf sheaths, and panicles. Plant

parts were carefully washed first with tap water and then with deionized water. After washing,

they were dried with paper and put into filter bags. Previously filter bags were weighed and

labelled. Bags with plant material were weighed to determine the fresh weight. Plant parts

were then dried until constant dry weight at 50°C. After drying the plant material, the bags

were weighed again to determine the dry weight. Plant parts were then ground and stored in

polyethylene beakers until analysis of heavy metal concentrations.


29

Figure 4.5. Common reed, Phragmites australis (original, 2004)

4.2.3. Sediment

Sediment was sampled at the inlet and outlet area of the presettlement tank and at the 6

different sampling positions along the length of the constructed wetland. A stainless steel

cylindrical corer was used to sample the sediment. Three replicate samples were taken at each

sampling position. Sediment was sampled into a cylindrical polyethylene core with a length of

30 cm. Sediment cores were packed into a polyethylene bags in order to ensure safe

transportation to the laboratory without altering sediment characteristics.

Figure 4.6. presents the sampled sediment cores. In the laboratory the sediment was divided

into 2 layers, an A layer with a depth of 0 to 15 cm below the sediment surface and a B layer

with a depth of 15 to 30 cm below the sediment surface. The A layer consisted of dark,

decomposing detrital material that was quite loosely structured. The B layer was darker than

the upper A layer and was dominated by greyish-black colour that is beneath vegetative mat.

Sediment samples of the presettlement tank were not distinguished into an A and B layer as

the sediment layer was shallow and the presence of concrete at the bottom of the

presettlement tank hampered the use of the sediment corer.

Figure 4.6. Sediment sampled in polyethylene cylindrical cores (original, 2004)


30

The redox potential of both layers of the sediment was determined immediately after opening

the sediment core and by pushing the redox electrode into the heart of the sediment. Sulphides

(AVS) in both layers of the sediment samples were analysed as soon as possible. Sediment

was dried until constant dry weight at 50 °C and then was ground. Sediment was stored in the

polyethylene bags. General sediment characteristics were determined: pH, electrical

conductivity (EC), cation exchange capacity (CEC), % of organic material (OM), and texture.

Total heavy metal concentrations in the sediment were determined. Exchangeable or

NH4OAc-extractable metal concentrations were determined together with the analysis of

CEC. Simultaneously extracted metals were determined together with the AVS analysis,

representing the potentially available metals in the sediment.

4.3. Analytical procedures

4.3.1. Wastewater

pH was determined using a pH electrode (HI 1230B plastic, double-junction, combination,

gel, Hanna Instruments, Temse, Belgium). Electrical conductivity was determined using an

EC electrode (Type WTW LF 537). Dissolved heavy metal concentrations were analysed

using ICP-OES (Varian Vista MPX, Varian, Palo Alto, CA) (Fig. 4.7.).

Figure 4.7. ICP –OES (original, 2004)


31

4.3.2. Phragmites australis

After weighing, the plant material was dried at 50°C until constant dry weight and then

weighed again. The percentage of dry weight was determined by dividing the dry weight with

the fresh weight.

0.5 g of dry plant material was weighed in a pyrex beaker of 100 ml. Samples were weighed

on an analytical balance. 10 ml of 65% HNO3 was added. The beaker was covered with a

watch glass and put on the plate at 130°C, 50 W for 1 hour. 4 ml of H2O2 in total was added in

aliquots of 1 ml. The samples were let to cool down. After this, filtration over blue ribbon

filter paper (MN 640 d Macherey-Nagel) was performed and the filtrate was collected in 50

ml flasks. The filter paper was rinsed with 1% HNO3 and the filtrate was diluted to the mark

with 1% HNO3. Flasks were closed with parafilm. After the destruction, total heavy metal

concentrations were analysed by ICP-OES (Varian Vista MPX, Varian, Palo Alto, CA). (Du

Laing et al., 2004)

4.3.3. Sediment

pH

10 g of dry sediment was weighed in a 100 ml pyrex beaker and 50 ml of deionized water was

added. The suspension was stirred manually. Samples were let to equilibrate for 18 h. The pH

of the supernatant was measured with a pH-electrode (HI 1230B plastic, double-junction,

combination, gel, Hanna Instruments, Temse, Belgium).

Electrical conductivity

Electrical conductivity of the sediment (EC) was determined by weighing 10 g of dry

sediment in a 100 ml pyrex beaker. 50 ml of deionized water was added. The suspension was

stirred for ½ h using the automatic stirrer and filtered over white ribbon filter paper (MN 640

m Macherey-Nagel). EC in the filtrate was determined using WTW LF 537 conductivity

meter (Wissenschaftlich Technischen Werkstäten, Weilheim, Germany).


32

Organic matter

Organic matter content was determined by ‘loss on ignition’. Exactly 1 g of dry sediment was

weighed in a crucible and the sediment was ashed at 550°C for 2 h. The empty crucible was

weighed, as well as the crucible containing the dry sediment. After ashing, the crucible and

ash were weighed in order to determine the organic matter content.

Texture

Texture analyses were performed at the “Instituut voor Bosbouw en Wildbeheer” (Institute for

Forestry and Game Management - IBW). Pretreatment of the sediment samples with H2O2

was performed to remove organic material, and with an acetate buffer solution to remove

CaCO3. Samples were rinsed three times by decantation after sedimentation. Samples were

stirred for 4 hours after addition of a dispersion solution. Texture of sediment samples was

determined by means of laser diffraction (Coulter LS200, Miami, FL). The clay fraction is

defined as the 0-6 µm fraction, whereas the sand fraction is the > 50 µm fraction.

Redox potential

Immediately after opening the polyethylene core that contained the sediment, the electrode

was pushed into the heart of the sediment, preventing air from leaking into the sediment. The

electrode is a combination of a Pt and a gel reference electrode (HI 3090 B/5) and the redox

potential was measured by connecting the electrode to a HI 9025 meter (Hanna Instruments,

Temse, Belgium). Measurements were corrected with regard to standard hydrogen electrode,

by calibrating in a solution of 0.033 M K3Fe(CN)6 en 0.033 M K4Fe(CN)6 in 0.1 M KCl.

Cation exchange capacity and exchangeable metals

The cation exchange capacity is determined by saturation of the adsorption complex with

ammonium, after which metals are analyzed in the percolate. In the percolation tube with

filter plate at the bottom, 2 spoons of sand were brought. Then, 5 g of sediment and 35 g of

cleaned sand were placed in the tube and covered with 10 g quartz-sand mixture. 150 ml of

1M NH4OAc, pH 7 was percolated through the percolation tube (Fig. 4.8). The percolate was

collected and exchangeable heavy metal concentrations were analyzed by ICP-OES (Varian

Vista MPX, Varian, Palo Alto, CA).


33

Figure 4.8. CEC experiment (original, 2005)

The excess of NH4+ was removed from the percolation tube by rinsing with 300 ml

denaturated ethanol 95% in fractions of 30-40 ml. The filtrate was discarded. The NH4+-

saturated sediment samples were then treated with 500 ml 1M KCl. The percolate was

collected into a 500 ml volumetric flask. After percolation, flasks were filled to the final

volume of 500 ml with 1M KCl. NH4+ in the percolate was determined by distillation in a

Kjeltec system (Kjeltec system II 1002 distilling unit). Hence, 50 ml of the percolate was

pipetted into a distillation flask where 20 ml of 2% boric acid indicator mixture (H3BO3) was

added together with 1 spoon of MgO. During distillation the change in color was observed

from red to greenish. The ammonia was titrated with 0.01 N HCl from green to blank color.

5 g of sediment and 500 ml KCl corresponds to 50 ml percolate and 0.5 g sediment.

Calculation of CEC in cmol kg-1 and in meq 100g-1 soil is as follows:

meq NH4+ g-1 = 2a x 0.01

meq NH4+ 100g sediment-1 = cmol (+) kg sediment-1 = 2a = CEC

a = ml 0.01 N HCl titrated

Acid volatile sulphide and simultaneously extracted metals

Acid volatile sulfide (AVS) is the most easily liberated sulphide fraction (Allen et al., 1995).

The principle of AVS determination is that first sulphides are transformed into H2S by

addition of concentrated HCl and then collecting the released H2S in a Zn acetate solution.

After addition of KIO3 and KI, the formed I2 reacts with the precipitated zinc sulphides (ZnS).

Sulphur and iodides are formed. The excess of I2 can be determined by back titration with

Na2S2O3 until color change with starch as an indicator. The difference between the added

quantity of KIO3 and the backtitrated quantity is a measure of the milli-equivalents of

sulphides.


34

KIO3 + 5KI + 6H2O 3I2 + 3H2O + 6KOH

I2 +S2- 2I- + S

I2 + 2 S2O32- 2I- + S4O6

2-

Determination of sulphides was performed with 3 airtight bottles filled with deionized water

that are connected to each other in series. The first bottle was connected to a gas bottle with

N2. Nitrogen gas passed through all 3 bottles before leaving the 3rd bottle. Before the

experiment was performed, O2 was removed from the bottles by passing N2 through them.

After 10 minutes, 5 ml ZnAc was added to the third and second bottle containing 95 ml of

deionized water. After that, 5 g of sediment and 20 ml of 6M HCl were added to 100 ml of

deionized water in the first bottle. The suspension was mixed by means of a magnetic stirrer.

After 5 minutes, all the bottles were subjected to a flux of N2 gas for 30 minutes. The

suspension from the first bottle was filtered into a 250 ml flask over a white ribbon filter (MN

640 m Macherey-Nagel) and diluted to the mark. The filtrate was analysed for heavy metals

by means of ICP-OES (Varian Vista MPX, Varian, Palo Alto, CA). The suspensions in the

other two bottles were quantitatively collected in an erlenmeyer after which 5 ml of 0.025 N

KIO3, 2 g of KI, 6 ml of concentrated HCl and 2 ml of starch solution (5 g L-1) as an indicator

were added. This was back titrated with 0.025 N Na2S2O3 solution until the solution was

colorless. A blank determination was performed.

Total heavy metal concentrations

The procedure of aqua regia destruction is based on transferring 1g of dry sediment (105°C)

into a pyrex beaker and adding 10 ml of aqua regia that consists of 7.5 ml concentrated

hydrochloric acid (HCl) and 2.5 ml of concentrated nitric acid (HNO3) (Fig. 4.9.). The work

was done under the fume hood. The container was covered with a watch glass and allowed to

react under the fume hood overnight, for a minimum of 12h. The following day samples were

put on the plate and boiled at 150°C, 50 W for 2 h. Samples were allowed to cool down to

ambient temperature. The extract was filtered over white ribbon filter (MN 640 6 Macherey-

Nagel) and collected into a 100 ml volumetric flask. The vessels and the residue on the filter

paper were rinsed with small amounts of 1% HNO3. The samples were diluted to 100 ml with

1% HNO3. Total heavy metal concentrations were determined by ICP-OES (Varian Vista

MPX, Varian, Palo Alto, CA).


35

Figure 4.9. Aqua Regia experiment (original, 2005)

4.4. Mass balance

4.4.1. Metal mass removed from wastewater

A rough estimation of the metal mass removed from the wastewater during the operational

lifetime of the constructed wetland was made. Metal concentrations were based on three

sampling times: October and November 2004, and August 2003. The latter data were used

from the thesis of I. Leuridan that worked on the same constructed wetland (Leuridan, 2004).

An important remark is that calculations only considered dissolved metal concentrations.

Metals that might be complexed and adsorbed to suspended solids were not taken into

account. This can lead to an underestimation of the real metal loading of the constructed

wetland. Flow data were obtained from Aquafin NV for the last 5 years of operation. The bed

has been in operation since 1989, but data is only available from the year the constructed

wetland was handed over to Aquafin NV. A mean flow rate of 508 ± 335 m³ d-1 was observed

during the last 5 years and was extrapolated to the entire 16 years of the operation. Formula

(1) was used to estimate the metal mass removed from the wastewater during 16 years of

operation:

Mwater = (Ceff – Cinf) * Q * t (1)

Where Mwater = Metal mass removed from wastewater (mg)

Cinf = Metal concentration in the influent (mg L-1)

Ceff = Metal concentration in the effluent (mg L-1)

Q = Daily flow (L d-1)

t = Time since the start of operation (d)

Due to the limited amount of the concentration data, the high spread on concentration levels

and flow data, and extrapolation errors, this mass should be considered as indicative, but not

very reliable.


36

4.4.2. Metal mass accumulated in aboveground Phragmites biomass

In order to estimate the contribution of the aboveground biomass to the total metal

accumulation in the constructed wetland and thus to estimate the metal masses in each plant

part, following theoretical assumptions were made:

• According to Gessner (2000) the total annual aboveground biomass production varies

from 1 to 2 kg dry mass m-2. A mean biomass production of 1.5 kg dry weight m-² was

used for the calculations.

• Stem and leaf sheaths occur together and are closely attached. When the total stem dry

mass is considered, leaf sheaths make up 25% and the stem material deprived of

sheaths makes up 75% of total stem dry mass (Gessner, 2001). Another experiment

from the same author (Gessner, 1996) described that stems together with sheaths

account for 75% of the total aboveground biomass production, leaves account for 24%

and panicles account for only 1% of the total aboveground biomass production.

Following procentual division of the aboveground biomass production can thus be

derived: stems, leaf sheaths, leaves, and panicles account for 56 %, 19 %, 24 %, and 1

% of the aboveground biomass production, respectively. Mean biomass production of

stems, leaf sheaths, leaves, and panicles are respectively 0.84 kg DW m-², 0.285 kg

DW m-², 0.36 kg DW m-², and 0.015 kg DW m-².

The total planted surface area of the 9 channels was 3060 m² (planted surface width of 4 m in

each channel). The total annual aboveground biomass production was estimated to be 4590 kg

dry weight of which 2570 kg, 872 kg, 1102 kg, and 46 kg was accounted for stems, leaf

sheaths, leaves, and panicles, respectively. The metal mass in each plant part was calculated

by formula (2). Summation of the metal masses in stems, leaf sheats, leaves, and panicles is

then the total metal mass accumulated in the aboveground biomass production.

Mplant part = Cplant part * DWplant part (2)

Where Mplant part = Metal mass accumulated in plant part (mg)

Cplant part = Metal concentration in plant part (mg kg-1)

DWplant part = Dry weight production of plant part (kg)

Plant part = Stems, leaf sheaths, leaves, or panicles


37

4.4.3. Metal mass accumulated in the sediment

The mean dry weight in a sediment core was determined (N = 18; mean = 516 ± 71 g). The

dry mass of the top 30 cm sediment layer in the reed bed was estimated by means of the ratio

of the surface area of one sediment core (Ø 4.5 cm, 0.00159 m²) and the total surface area of

the 9 channels (1530 m², bottom width 2 m). The top sediment layer in the entire reed bed had

a total dry mass of 500 ton DW. The metal mass accumulated in the sediment was calculated

by formula (3):

Msediment = Csediment * DWsediment (3)

Where Msediment = Metal mass in sediment (mg)

Csediment = Metal concentration in sediment (mg kg-1)

DWsediment = Dry weight of sediment (kg)

4.5. Detection limits

Detection limits of ICP-OES were determined by the sum of the mean of blank values and

twice the standard deviation of the blanks (adopted from Wustenbergs, 2004) (Table 4.2).

Table 4.2. Detection limits of ICP-OES expressed in µg L-1

Metal Detection limit Metal Detection limit

Cd 1 Ni 5

Cr 5 Zn 15

Cu 10 Fe 35

Pb 10 Mn 10

Results

38

5. RESULTS

5.1. Wastewater

5.1.1. pH and electrical conductivity (EC)

The pH of the wastewater was unaffected by sampling position along the treatment path of the

SF wetland. The pH of the wastewater sampled in October had a mean value of 7.8 ± 0.2,

while the wastewater sampled in November had a mean pH of 7.6 ± 0.2. The pH in the

surface flow reed bed had a mean value of 7.7 ± 0.1.

The electrical conductivity (EC) of the water samples was 872 ± 40 µS cm-1 during the first

sampling. During the second sampling the wastewater had an EC of 976 ± 11 µS cm-1 (Fig.

5.1.). A more or less constant EC of 900 - 1000 µS cm-1 was described in the SF wetland.

0

200

400

600

800

1000

1200

1400

inf eff PT 0 25 70 140 350 630 effDistance (m)

EC (µ

S cm

-1)

1st water sampling

2nd water sampling

Figure 5.1. Electrical conductivity as a function of sampling position in the constructed

wetland

inf : influent of the presettlement tank; eff PT : effluent of the presettlement tank; eff: effluent

of SF wetland; 0, 25, 70, 140, 350, 630: distance in meters from the inlet of the SF

The influent of the presettlement tank denotes the influent of the entire wastewater treatment

system and the effluent of the presettlement tank can be considered as the influent of the SF

reed bed.

Results

39

5.1.2. Dissolved heavy metal concentrations

Generally, dissolved metal concentrations in the wastewater along the treatment path in the

SF wetland were low. During the first sampling period (October 2004) Al, Cd, Cr, Cu, Fe,

Mn, Ni, Pb, and Zn were detected in the wastewater. Five of these elements: Al, Cu, Fe, Mn,

and Zn were detected during the second sampling (November 2004) as well, whereas Cd, Cr,

Ni, and Pb concentrations could not be detected. Dissolved concentrations of all metals in the

wastewater did not exceed the surface water quality criteria described in Vlarem II.

Dissolved concentrations of Al, Cu, and Zn in the wastewater clearly showed a decreasing

trend as a function of increasing distance from the inlet of the reed bed (Fig. 5.2.). The Al

concentration in the influent of the system (90 ± 3 µg l-1) decreased to 64 ± 40 µg l-1 in the

reed bed effluent during the first sampling in October 2004.

During the second sampling period a similar Al concentration was seen in the influent of the

system (109 ± 74 µg l-1), but the concentration in the effluent was lower (14 ± 3 µg l-1) and

thus a higher removal efficiency occurred. As for Al a decrease in Cu and Zn concentrations

was seen in the wastewater. Cu concentrations decreased from 14 ± 6 µg l-1 and 19 ± 4 µg l-1

in the influent to 5 ± 2 µg l-1 and 4 ± 1 µg l-1 in the effluent at both sampling dates. Influent Zn

concentrations in November 2004 (129 ± 12 µg l-1) were higher than in October 2004 (75 ±

27 µg l-1), but similar effluent concentrations were seen at both sampling dates (32 ± 10 µg l-1

and 34 ± 5 µg l-1 in October and November 2004, respectively).

Results

40

0

20

40

60

80

100

120

140

160

180

200

inf ef f PT 0 25 70 140 350 630 eff

Distance (m)

Al 1st samplingAl 2nd sampling

A

0

5

10

15

20

25

30

inf eff PT 0 25 70 140 350 630 eff

Distance (m)

Cu

(µg

l-1)

Cu 1st samplingCu 2nd sampling

B

0

20

40

60

80

100

120

140

160

inf ef f PT 0 25 70 140 350 630 eff

Distance (m)

Zn 1st samplingZn 2nd sampling

C Figure 5.2. Dissolved Al (A), Cu (B), and Zn (C) concentration in the wastewater as a

function of sampling position expressed in µg l-1

Results

41

The dissolved Fe concentration ranged from 29 ± 5 µg l-1 to 163 ± 67 µg l-1 during the first

sampling period whereas the spread on the values was lower in November 2004 (Fig. 5.3.A).

Dissolved Fe concentrations in the effluent were lower than in the influent at both sampling

dates. A different trend was described for Mn. An increase of dissolved Mn concentrations

along the treatment path of the SF wetland was seen at both sampling dates. Influent

concentrations (49 ± 3 µg l-1 and 40 ± 2 µg l-1 in October and November 2004, respectively)

increased to 76 ± 1 µg l-1 and 71 ± 0.27 µg l-1 in the effluent of the reed bed (Fig. 5.3.B).

0

50

100

150

200

250

inf eff PT 0 25 70 140 350 630 eff

Distance (m)

Fe (µ

g l-1

)

1st sampling2nd sampling

A

0

10

20

30

40

50

60

70

80

90

inf eff PT 0 25 70 140 350 630 eff

Distance (m)

Mn

(µg

l-1)

1st sampling 2nd sampling

B Figure 5.3. Dissolved Fe (A), and Mn (B) concentration in the wastewater as a function of

sampling position expressed in µg l-1

Cd, Cr, Ni, and Pb were present in very small concentrations at both sampling dates and were

below detection limits in November 2004. Mean dissolved concentrations of Cd, Cr, Ni, and

Pb in the wastewater in October 2004 are presented in Table 5.1. For this group of metals no

decreasing or increasing trends were observed along the treatment path of the CW, except for

Pb that showed a slight decrease with increasing distance from the inlet.

Results

42

Table 5.1. Dissolved Cd, Cr, Ni, and Pb concentration in the wastewater in October 2004

expressed in µg l-1

Metal Dissolved concentration (µg l-1)

Cd 0.55 ± 0.07

Cr 1.01 ± 0.09

Ni 3 ± 0.45

Pb 7 ± 2

5.1.3. Removal efficiency of the SF wetland

Mean removal efficiencies at both sampling dates are presented in Table 5.2. The removal

efficiency of the surface flow wetland varied between metals. Al, Cu, and Zn were efficiently

removed at both sampling dates. Removal efficiencies of Cd, Cr, Ni, and Pb need to be

considered carefully as concentrations levels in the wastewater were very low. A negative

removal efficiency (as for Cd, Cr and Ni in October 2004) does not indicate per se that Cd,

Cr, and Ni were released from the wetland. Metal concentrations in the influent wastewater

were already very low and these metals were present at a certain background concentration.

Negative removal efficiencies of Mn on the other hand and the concentration levels in the

wastewater as described in the previous paragraph, do indicate that Mn was released from the

wetland at this stage of operation.

Table 5.2. Mean removal efficiencies (%) of metals in October and November 2004

Metal Removal efficiency (%), Oct 2004 Removal efficiency (%), Nov 2004

Al 28 63

Cd -5 -

Cr -12 -

Cu 64 75

Fe 32 -71

Mn -55 -103

Ni -50 -

Pb 40 -

Zn 57 62

Results

43

5.1.4. Metal mass removed from the wastewater

Taking into account the concentration levels described in paragraph 5.1.2. and a mean flow of

508 ± 335 m³ d-1 during 16 years of operation, the metal mass removed from the wastewater

since the start of operation was estimated and is presented in Table 5.3. After 16 years of

operation 114 ± 121 kg of Zn and 27 ± 19 kg of Cu would have been removed from the

wastewater and would have accumulated in the SF wetland.

Table 5.3. Metal mass removed from the wastewater expressed in kg

Metal Mass removed from the wastewater (kg)

Cr 2± 3

Cu 27± 19

Cd 2± 2

Ni 1± 4

Pb 18± 26

Zn 114± 121

Results

44

5.2. Phragmites australis biomass

Phragmites australis biomass was differentiated into four different plant parts and the highest

% of dry weight was determined in leaves (Table 5.4.). Metal concentrations in different plant

parts and the metal mass accumulated in the aboveground biomass are presented and

discussed in paragraphs 5.2.1. and 5.2.2.

Table 5.4. % dry weight in different plant parts

Plant part % dry weight

Leaves 48 ± 6

Stems 35 ± 4

Panicles 28 ± 4

Leaf sheats 38 ± 10

5.2.1. Metal concentrations in Phragmites australis biomass

Metal concentrations in different plant parts of Phragmites australis as a function of distance

from the inlet are presented in Appendix I. Metal concentrations in aboveground biomass

appeared to be unaffected by sampling position in the surface flow wetland and showed

normal variations. Mean metal concentrations in the aboveground plant parts are presented in

Table 5.5.

The highest concentrations of Al, Cu, Ni, Pb, and Zn were seen in the panicles of Phragmites

australis. Fe and Mn had the highest concentrations in the leaves. The Cr concentration in

stems was the highest, compared to the other plant parts. The lowest uptake by all four plant

parts was for Cd ranging from 0.03 ± 0.01 to 0.05 ± 0.01 mg kg-1, whereas Mn showed the

highest uptake of all metals in the leaves (240 ± 108 mg kg-1). Cd, Cr and Ni were present in

very low concentrations in the aboveground biomass.

Results

45

Table 5.5. Metal concentrations in different plant parts of Phragmites australis in mg kg-1DW

Metal Stems Leaves Leaf sheets Panicles

Al 8 ± 1.4 18± 4 13± 3 34 ± 26

Cd 0.03 ± 0.01 0.05± 0.01 0.05± 0.01 0.05 ± 0.01

Cr 1.32 ± 0.37 0.81± 0.14 0.82± 0.24 0.96 ± 0.29

Cu 1.34 ± 0.41 4± 0.93 1.92± 0.4 5 ± 1.97

Fe 34 ± 7 153± 17 55± 9 144 ± 75

Mn 30 ± 12 240± 108 88± 29 80 ± 15

Ni 0.5 ± 0.17 0.7± 0.17 0.8± 0.15 0.81 ± 0.22

Pb 0.45 ± 0.2 1.07± 0.24 0.71± 0.19 1.73 ± 1.05

Zn 72 ± 30 51± 10 38± 6 82 ± 15

The observed metal concentrations in Phragmites australis plant parts of the SF wetland in

Deurle can be compared to results reported by other authors (Table 5.6.). For example, the Cd

concentration in aboveground parts of reed ranged from 0.03 ± 0.01 mg kg-1 in stems to 0.05

± 0.01 in leaves, leaf sheaths and panicles. These Cd concentrations are comparable to

concentrations found in aboveground plant parts in 2 lakes in Denmark (Schierup & Larsen,

1981; Vymazal, 2003) and to concentrations found in a constructed wetland that treats landfill

leachate (Surface et al., 1993). The Fe concentration ranged from 33.6 ± 7.4 mg kg-1 in stems

to 152 ± 17 mg kg-1 in leaves, concentrations which are quite low compared to the values

reported in Table 5.6. Fe concentrations are comparable with those in a constructed wetland in

the Czech Republic (Zuidervaart, 1996; Vymazal, 2003). Lowest Ni concentrations are

encountered in the stems (0.49 ± 0.16 mg kg-1) whereas highest concentrations are

encountered in the panicles (0.81 ± 0.22 mg kg-1). The observed Ni concentrations are lower

than in Tanzania where they vary between 1.3-9.2 mg kg-1 (Ojo & Mashauri, 1996). The

observed Pb concentrations in aboveground plant parts are comparable with those in plants of

constructed wetlands in the Czech Republic (Zuidervaart, 1996; Vymazal, 2003). The highest

Pb concentration was encountered in the panicles (1.73 ± 1.05 mg kg-1). Mn concentrations in

aboveground plant parts were comparable to those reported by Ye et al. (2001) in the study of

a 10-year-old CW treating leachate (312 ± 38 mg kg-1 Mn). The same author showed that

other metals except Mn, had the highest concentrations in the roots and that those were

comparable to the concentrations in fallen litter.

Results

46

Table 5.6. Concentration of some selected metals in Phragmites australis aboveground plant

parts found in different constructed wetlands expressed in mg kg-1

Locality Type Plant part Cd Fe Ni Pb Ref.

Czech Rep. Eutrophic

pond

AB - 152-2773 - 2.6-5.4 1

Czech Rep. 4 HSF CW’s AB 3.6-9.5 64-1341 1.3-1.9 1

AB 173-215 1343-1398 - - 2 India VF CW

R 1699-1782 12743-12942 - 2

Bulgaria NW AMD AB 7-37 820-1450 - - 3

Tanzania Various WW R 1.3 7905-18850 1.3-9.2 10-11.3 4

Poland Lake AB - - - 2.1 5

UK NW AB - - - 264 6

Denmark 2 lakes AB <0.1-2.5 - - 0.05-3.7 7

New York CW LL AB 0.09 66 - 0.21 8

AB=aboveground biomass, R=roots, CW=constructed wetland, NW=natural wetland, AMD=

acid mine drainage, LL=landfill leachate, WW-wastewater

1-Zuidervaart (1996); 2–Oke & Juwarkar (1996); 3-Groudev et al. (2002); 4-Ojo & Mashauri,

(1996); 5-Kufel (1991); 6-Mungur et al. (1994); 7-Schierup & Larsen (1981); 8-Surface et al.

(1993)

5.2.2. Metal mass accumulated in aboveground Phragmites biomass

The metal mass accumulated in the aboveground biomass is presented in table 5.7. At this

moment there is a lack of data of the aboveground biomass production of Phragmites

australis in the SF wetland in Deurle. The mass calculations were therefore based on an

estimation of the aboveground biomass production, as described in chapter 3. It must

therefore be considered as indicative and not as an absolute measure of the metal mass

accumulated in the aboveground plant parts. The highest masses were encountered for the

essential micronutrients Mn, Fe, and Zn (0.42 ± 0.13, 0.31 ± 0.028, and 0.28 ± 0.079 kg,

respectively). Masses of trace metals Cr, Cd, Ni, and Pb were very low.

Results

47

Table 5.7. Metal mass accumulated in aboveground Phragmites biomass expressed in kg

Metal Metal mass (kg)

Cr 0.0051 ± 0.0010

Cu 0.0094 ± 0.0015

Cd 0.0002 ± 0.00003

Fe 0.31 ± 0.028

Mn 0.42 ± 0.13

Ni 0.0027 ± 0.0005

Pb 0.0030 ± 0.0006

Zn 0.28 ± 0.079

Results

48

5.3. Sediment

5.3.1. Sediment characteristics

5.3.1.1. General sediment characteristics

No clear spatial effect was observed for most sediment characteristics (pH, EC, % OM, %

DW, CEC, texture, and exchangeable concentrations of the main elements). Table 5.8.

presents the sediment characteristics, with a distinction between the sediment sampled in the

presettlement tank and the two sediment layers in the SF reed bed (layer A: 0-15 cm; layer B:

15-30 cm). Texture analysis was performed on composite sediment samples at each sampling

position, without a distinction in depth. The mean clay, sand, and silt % of the sediment was

15 ± 4, 53 ± 7, and 31 ± 5 % respectively. The 6 sampling positions in the reed bed had

different textures, varying from light clay, light sand silt to heavy sand silt.

Table 5.8. Sediment characteristics of the presettlement tank and the surface flow reed bed

Parameter Presettlement tank Layer A (0-15 cm) Layer B (15-30 cm)

EC (µS cm-1) 351 ± 12 381 ± 147 262 ± 31

pH 7.79 ± 0.05 7.67 ± 0.1 7.76 ± 0.06

OM (%) 3.5 ± 0.14 5.93 ± 3 6 ± 2

DW (%) 68 ± 2 63 ± 11 70 ± 4

CEC (cmol (+) kg-1DW) 5 ± 1 7 ± 2 9 ± 4

Ca (cmol (+) kg-1DW) 38 ± 6 47 ± 9 53 ± 12

K (cmol (+) kg-1DW) 0.19 ± 0.05 0.35 ± 0.09 0.47 ± 0.24

Na (cmol (+) kg-1DW) 0.4 ± 0.13 0.48 ± 0.26 0.48 ± 0.22

Mg (cmol (+) kg-1DW) 0.7 ± 0.23 1.13 ± 0.33 1.25 ± 0.53

The sediment pH varied from 7.76 to 7.79, a value corresponding to the pH of 7.0 ± 0.1

reported by Leuridan (2004). The electrical conductivity of the deeper sediment layer (262 ±

31 µS cm-1) had a lower value than the upper 15 cm sediment layer and the sediment in the

presettlement tank (respectively 381 ± 147 µS cm-1 and 351 ± 12 µS cm-1). The electrical

conductivity reported by Leuridan (2004) (270 ± 12 µS cm-1) resembles the observed EC in

the deeper sediment layer.

Results

49

The cation exchange capacity of the sediment of the presettlement tank (5 ± 1 cmol (+) kg-1

DW) was lower than that of the sediment of the SF reed bed (Table 5.8.). The CEC of the

sediment of the SF reed bed was however lower than the value of 12 ± 1 cmol (+) kg-1 DW

reported by Leuridan (2004). The % of organic matter of the sediment in the SF reed bed

was about 6 %, which is slightly less than the 7.6 ± 0.5 % reported by Leuridan (2004). The

organic matter content of the sediment in the presettlement tank is remarkably lower (3.5 ±

0.14 %). A reason for the higher organic matter content and cation exchange capacity of the

sediment in the SF reed bed could be the input of decomposing Phragmites australis biomass.

Indeed, if reeds are not harvested a large amount of organic material is eventually cycled to

the sediment, thereby increasing the CEC and the amount of organic matter. A standard soil

has an organic matter content of 2 %, which is considerably lower than that of the sediment of

the SF reed bed.

The exchangeable Ca concentration was the highest for all main elements. However, it has to

be noted that the exchangeable Ca concentration exceeds the CEC of the sediment, indicating

that other forms of Ca were extracted by the 1M NH4OAc reagent as well. If Ca would be

excluded from the comparison, then following metals dominated in the following order Mg >

Na > K. Ca, Mg, and K are essential plant nutrients and can be readily exchanged from the

negatively charged sediment surface sites.

5.3.1.2. Redox potential

The redox potential of the sediment of the presettlement tank and the upper sediment layer (0-

15 cm) of the SF reed bed was unaffected by sampling position (Fig. 5.4.). A mean redox

potential of –129 ± 60 mV was seen in the presettlement tank, a result similar to the mean

redox potential in the top sediment layer of the SF reed bed (-138 ± 42 mV). The redox

potential of the deeper sediment layer (15-30 cm) appeared to be affected by sampling

position with lower values in the first 25 m of the wetland (Fig. 5.4.). This indicates very

reduced conditions at the inlet area of the reed bed, which might be explained by the higher

organic loading at the inlet area.

Results

50

-450

-400

-350

-300

-250

-200

-150

-100

-50

0inf eff PT 0 25 70 140 350 630

Distance (m)

Red

ox p

oten

tial (

mV)

0-15 cm15-30 cm

Figure 5.4. Redox potential of the sediment of the presettlement tank and the upper and

deeper sediment layer in the SF reed bed, expressed in mV

5.3.1.3. Acid volatile sulphide (AVS)

Acid volatile sulphide (AVS) is a measure for the amount of metal sulphides in the sediment

and is presented in Fig. 5.5. A clear spatial effect was not seen in the deeper sediment layer,

whereas a decrease of the AVS concentration could be noted in the upper sediment layer with

increasing distance from the inlet area. Higher AVS values in the first 25 m of the reed bed

could be caused by the higher organic loading in the wastewater at the inlet area. As lower

redox potentials were seen in the deeper sediment of the first 25 m of the reed bed (Fig. 5.4.),

one would expect higher AVS concentrations in the deeper sediment layer as well. This is not

the case. Sulphides can only be formed in very reduced sediments in the presence of organic

matter. However, both sediment layers have equal organic matter contents (about 6 %, Table

5.8.), which can therefore not explain the higher sulphides content in the inlet area of the

upper sediment layer. Further along the treatment path of the SF wetland, AVS levels in both

sediment layers are more or less alike.

The low AVS level of the deeper sediment layer at 140 m from the inlet of the SF reed bed

does not comply with the other AVS levels in the deep sediment and is considered unreliable.

Oxidation of the sediment samples during storage and handling of the samples could have led

to the low AVS level at this sampling position.

Results

51

-1000

0

1000

2000

3000

4000

5000

6000

7000

8000

inf PT ef f PT 0 25 70 140 350 630

Distance (m)

mg

S2- k

g-1 D

M

0-15 cm

15-30 cm

Figure 5.5. AVS of the sediment of the presettlement tank and the upper and deeper sediment

layer in the SF reed bed, expressed in mg S2- kg-1 DM

5.3.2. Metal concentrations in the sediment

5.3.2.1. Total metal concentrations in the sediment

Total metal concentrations in the sediment are presented in Table 5.9. Stratification of the

sediment of the presettlement tank was not possible as the sediment layer was too shallow.

Four important conclusions were drawn:

1) A spatial effect on the total metal concentrations in the sediment of the SF reed bed

was not seen for Pb, Mn, Fe, Cd, and Cr. Total concentrations in both sediment layers

showed normal variations without a specific increasing or decreasing trend. The total Zn, Ni,

and Cu concentration in the upper sediment layer (0 – 15 cm) at the inlet area of the SF reed

bed was higher than at the other sampling positions further along the treatment path.

However, high standard deviations are present at the first sampling position, making it

difficult to draw straightforward conclusions about spatial effects on total Zn, Ni, and Cu

concentrations. Total metal concentrations in both sediment layers of the SF reed bed as a

function of distance from the inlet are presented in Appendix II. Mean metal concentrations in

both sediment layers of the SF reed bed are presented in Table 5.9.

Results

52

2) Metal concentrations in the sediment sampled in the outlet area of the presettlement

tank were higher than those in the sediment sampled in the inlet area. This was seen for

all metals under study and is somewhat contradictory to what one would expect. Under

normal conditions, metals present in a particulate form in the wastewater are thought to settle

out in the inlet area of the presettlement tank. The presence of an overflow construction

before the presettlement tank allows to discharge wastewater directly into the surface water in

case of storm conditions. However, it can be assumed that with the regular occurrence of

storm weather, sediment could be flushed out of the inlet area and settle again at the outlet

area, leading to elevated metal concentrations in this area.

3) Stratification of total metal concentrations in the sediment of the SF reed bed

occurred for all metals except Ni. The deeper sediment layer (15 – 30 cm) generally had

higher total metal concentrations than the upper sediment layer. However, Ye et al. (2001)

described higher Fe, Cd and Zn concentrations in the top layer (0 – 5 cm) of a CW that treats

coal ash pile leachate.

4) Total metal concentrations in the sediment of the SF reed bed and in the sediment of

the outlet area of the presettlement tank were elevated compared to background values

but were lower than the soil remediation criteria (Vlarebo, 1996). The pollution level of

the sediment is not considered to be that elevated to form an immediate risk. Total Cd, Cr,

Cu, Ni, and Pb concentrations in the sediment of the inlet area of the presettlement tank were

lower than or similar to the background values.

Results

53

Table 5.9. Total metal concentrations in the sediment of the presettlement tank and both

sediment layers of the SF reed bed, with background values and soil remediation standards

corrected for % OM and % clay, all expressed in mg kg-1 (Vlarebo, 1996)

Metal Inf PT Eff PT 0 –15 cm 15-30 cm Background

values

Soil remediation

standards

Cd 0.28± 0.12 1.04 ± 0.38 1.4± 0.2 2.1±0.7 1.15 2.88

Cr 23± 3 45 ± 13 48± 3 67±5 40 141

Cu 25± 7 46 ± 10 60± 36 69±8 19 218

Fe 5893± 425 9350 ± 1576 8930± 757 10630±1208 - -

Mn 64± 11 96 ± 16 120± 5 157±11 - -

Ni 11± 2 16 ± 3 19± 1.6 19±2 11 124

Pb 36± 4 80 ± 32 96± 13 150±25 51 253

Zn 140± 48 260 ± 58 318± 163 386±90 77 743

Inf PT and Eff PT represent the inlet and outlet area of the presettlement tank respectively

The total metal concentrations in the sediment had following order:

Fe > Zn > Mn > Pb > Cu > Cr > Ni > Cd

The total concentrations of Fe and Zn in the sediment were respectively 9000 – 10000 mg kg-1

and 320 – 380 mg kg-1, being the highest among all metals. Ye et al. (2001) reported Fe

concentrations of 70 000 mg kg-1 in the sediment of a CW treating metal contaminated

leachate from a coal ash pile. Fe and Mn concentrations of 32 300 and 1200 mg kg-1 were

reported in the top 15 cm of the sediment of a CW treating leachate from an electrical power

station. Vymazal (2003) reported high concentrations of Cd and Ni (27.5 and 45.4 mg kg-1

respectively) in the sediment of a CW receiving municipal wastewater. Cd and Ni

concentrations in the sediment of the CW in Deurle were much lower. The Pb concentration

in the sediment of the CW in Deurle was comparable to the value of 155 mg kg-1 reported by

Vymazal (2003). Sawidis et al. (1995) reported sediment concentrations of Cu, Zn, Ni, Cd,

Pb, and Mn between 19.5 - 27.6, 42.5 - 95, 32 - 230, 1.1 - 3.3, 6.5 - 20.5, 424 - 1000 mg kg-1

in aquatic systems receiving sewage, respectively. Cu and Cd concentrations in the sediment

of this study were similar, whereas Zn and Pb concentrations were elevated. Ni and Mn

concentrations in the sediment of the SF wetland in Deurle were considerably lower.

Results

54

5.3.2.2. Exchangeable metal concentrations

Table 5.10. presents exchangeable metal concentrations in both sediment layers of the SF reed

bed and in the sediment in the presettlement tank. Exchangeable metal concentrations

appeared to follow similar trends as the total metal concentrations:

• Exchangeable metal concentrations in the sediment of the SF reed bed were unaffected

by sampling position along the treatment path. Mean exchangeable metal

concentrations are therefore presented in Table 5.10.

• Exchangeable metal concentrations in the sediment sampled in the outlet area of the

presettlement tank were higher than those in the sediment sampled in the inlet area.

This was seen for all metals and is similar to what is observed for the total metal

concentration (5.3.2.1.).

• A stratification of the exchangeable metal concentrations in the sediment of the SF

reed bed occurred for all metals except Ni, where the deeper sediment layer (15 – 30

cm) generally had higher exchangeable metal concentrations than the upper sediment

layer.

Table 5.10. Exchangeable metal concentrations in the sediment of the presettlement tank

and both sediment layers of the SF reed bed expressed in mg kg-1

Metal Inf PT Eff PT 0 - 15 cm 15-30 cm

Al 0.18± 0.12 0.35± 0.13 0.25± 0.08 0.53± 0.09

Cd 0.11± 0.04 0.16± 0.05 0.24± 0.07 0.41± 0.37

Cr 0.07± 0.03 0.13± 0.06 0.11± 0.03 0.15± 0.06

Cu 0.50± 0.06 0.62± 0.09 0.99± 0.45 1.69± 1.06

Fe 0.35± 0.16 0.61± 0.18 0.50± 0.12 0.76± .0.25

Mn 9± 2 12± 3 16± 3 19± 4

Ni 0.52± 0.18 0.77± 0.2 0.66± 0.16 0.56± 0.24

Pb 1.46± 0.3 2.39± 0.7 3± 1 5± 4

Zn 14± 4 22± 6 23± 4 25± 14

Inf PT and Eff PT represent the inlet and outlet area of the presettlement tank, respectively

Results

55

When the exchangeable metal concentrations in the SF reed bed are compared to their total

concentrations (Table 5.11.) following order of decreasing exchangeability is noted:

Cd > Mn > Zn > Pb ~ Ni > Cu > Cr > Fe

Cd, Mn, and Zn appeared to be the most mobile metals in the sediment with 15 – 39, 12-14,

and 6-10 % of the total metal content being exchangeable, respectively. These elements are

considered to be the most mobile ones.

Table 5.11. % of exchangeable metals as a function of sediment depth

Metal Inf PT Eff PT 0 - 15 cm 15-30 cm

Cd 39 15 17 20

Cr 0.3 0.3 0.23 0.22

Cu 2 1 2 2

Fe 0.01 0.01 0.0056 0.0072

Mn 14 13 14 12

Ni 5 5 3 3

Pb 4 3 4 4

Zn 10 8 7 6 5.3.2.3. SEM - simultaneously extracted metals

Metal sulphides are present in the sediment (5.3.1.3.). Together with the analysis of acid

volatile sulphides (AVS), the simultaneously extracted metals were determined as well and

mean values in both sediment layers of the SF reed bed are presented in Table 5.12. The SEM

values at different sampling positions in the constructed wetland showed variations without a

clear trend and are presented in Appendix III. Simultaneously extracted concentrations of Cd,

Cr, Cu, Fe, Mn, Ni, Pb, and Zn were slightly higher in the deeper sediment layer.

Simultaneously extracted concentrations of Cd, Mn, Pb, and Zn are comparable to their total

concentrations in the sediment (Table 5.9.).

Results

56

Table 5.12. Simultaneously extracted metal concentrations in both sediment layers of the SF

reed bed expressed in mg kg-1

Metal 0 –15 cm 15-30 cm

Al 1416 ± 1064 1285 ± 469

Cd 1 ± 0.4 2 ± 1.1

Cr 13 ± 3 21 ± 11

Cu 4 ± 7 18 ± 22

Fe 4149 ± 1878 4890 ± 1978

Mn 96 ± 24 129 ± 45

Ni 8 ± 2 10 ± 5

Pb 90 ± 31 125 ± 76

Zn 270 ± 170 390 ± 201

5.3.3. SEM/AVS ratio

Table 5.13. presents AVS levels and SEM/AVS ratio in both sediment layers of the SF reed

bed as a function of distance to the inlet. AVS represents the amount of sulphides in the

sediment available for binding metals and is a major parameter with respect to toxicity

prediction of heavy metals in anaerobic sediments (Van den Hoop et al., 1997). SEM

represents the amount of metals in the sediment that could be potentially available. If SEM

exceeds AVS, the sediments are potentially toxic (Di Toro et al., 1990; Hansen et al., 1996).

The SEM/AVS ratio could thus be used for toxicity assessment. When SEM/AVS ratio is < 1,

there may be no acute toxicity for aquatic organisms in terms of heavy metals. On the other

hand, when SEM/AVS > 1 the sediment may be considered potentially toxic (Fang et al,

2005). Metal sulphides have a very low solubility and due to that these metals are not

available for uptake by organisms if an excess of available sulphides is present.

Similar SEM/AVS ratios were reported in both sediment layers at each sampling position,

although slightly higher ratios were seen in the deeper sediment layer at 25, 70, and 630 m

from the inlet of the SF reed bed (Table 5.13.). The very high SEM/AVS ratio of 2.5 in the

deep sediment layer at 140 m from the inlet is attributed to a low AVS level at this position

and is considered unreliable. The low AVS level of 4 µmol g-1 does not comply with the other

Results

57

AVS levels in the deep sediment varying between 26 and 59 µmol g-1 and is considered

unreliable.

SEM/AVS ratios in the upper sediment layer increased with increasing distance from

the inlet. This is mainly attributed to the fact that AVS levels are higher at the inlet area of

the SF reed bed and decrease with increasing distance. The decrease of sulphides available to

precipitate and immobilize metals, leads to a higher potential availability of metals further

along the treatment path of the wetland. So although total metal levels are more or less

constant in the upper sediment layer (paragraph 5.3.2.1.) the potential availability is different.

With the exception of the unreliable SEM/AVS ratio in the deep sediment at 140 m from the

inlet, all SEM/AVS ratios were considerably lower than 1. The SEM/AVS ratio had a

maximum value of 0.4 in the deep sediment at 70 m from the inlet. The SEM/AVS ratios

indicate that metals are precipitated as sulphides and are not potentially available.

Table 5.13. AVS levels (in µmol g-1) and SEM/AVS ratio in both sediment layers as a

function of distance from the inlet

Distance AVS SEM/AVS

(m) Cu Cu+Pb Cu+Pb+Cd Cu+Pb+Cd+Ni Cu+Pb+Cd+Ni+Zn

0-15 cm

0 139 0.0000 0.004 0.000 0.01 0.07

25 91 0.0001 0.006 0.01 0.01 0.05

70 38 0.0006 0.008 0.01 0.01 0.08

140 58 0.0000 0.011 0.01 0.01 0.07

350 23 0.0001 0.016 0.02 0.02 0.16

630 15 0.0160 0.035 0.04 0.04 0.21

15-30 cm

0 59 0.0002 0.004 0.00 0.01 0.04

25 48 0.0004 0.009 0.01 0.01 0.09

70 26 0.0109 0.045 0.05 0.06 0.40

140 4 0.2107 0.487 0.49 0.54 2.50

350 40 0.0005 0.010 0.01 0.02 0.12

630 36 0.0112 0.024 0.02 0.03 0.27

Results

58

5.3.4. Metal mass accumulated in the sediment

Based on the total metal concentrations in the sediment, an estimation of the metal mass

accumulated in the sediment was made and is presented in Table 5.14. About 30 kg of both Cr

and Cu was accumulated in the top 30 cm sediment layer after 16 years of operation of the

CW. The Cd mass was low whereas the accumulated Zn mass was 175 ± 34 kg.

Table 5.14. Metal mass accumulated in the sediment expressed in kg

Metal Metal mass

Cr 28 ± 8

Cu 32 ± 5

Cd 1 ± 0.27

Fe 4850 ± 893

Mn 69 ± 16

Ni 10 ± 1

Pb 61 ± 21

Zn 175 ± 34

Discussion

59

6. DISCUSSION

6.1. Metal accumulation in the constructed wetland

6.1.1. Relative importance of sediment and Phragmites biomass in total metal

accumulation in the SF wetland

Metal masses in different compartments of the CW were estimated and are presented in Table

6.1. The metal mass removed from the wastewater during 16 years of operation is only a

rough estimation as it is based on 3 wastewater-sampling dates. Regular monitoring of

dissolved, total metal concentrations and flow rate would allow making more accurate

estimations of the metal mass removed from the wastewater. Although the calculation was

rough and based on several assumptions, the Cu, Cd, and Zn mass removed from the

wastewater had the same order of magnitude as their mass accumulated in the sediment

(Table 6.1.). According to the Cr, Ni, and Pb mass accumulated in the sediment of the reed

bed (28 ± 8 kg, 10 ± 1 kg, 61 ± 21 kg, respectively), their mass removed from the wastewater

was underestimated (2 ± 3 kg, 1 ± 4 kg, 18 ± 26 kg, respectively). The highest masses

removed from the wastewater during the operational lifetime of 16 years were of Zn and Cu

(114 ± 121 and 27 ± 19 kg, respectively), which could be caused by their higher abundance in

domestic wastewater.

Table 6.1. Metal mass removed from the wastewater and metal mass accumulated in the

sediment and Phragmites australis biomass, expressed in kg

Metal Wastewater Sediment Phragmites biomass

Cr 2± 3 28± 8 0.0051 ± 0.0010

Cu 27± 19 32± 5 0.0094 ± 0.0015

Cd 2± 2 1± 0.27 0.0002 ± 0.00003

Fe - ± - 4850± 893 0.31 ± 0.028

Mn - ± - 69± 16 0.42 ± 0.13

Ni 1± 4 10± 1 0.0027 ± 0.0005

Pb 18± 26 61± 21 0.0030 ± 0.0006

Zn 114± 121 175± 34 0.28 ± 0.079

Discussion

60

The metal mass accumulated in the CW is mainly present in the sediment. For all metals more

than 99.99 % of their mass is accumulated in the sediment. Accumulation of metals in

aboveground plant parts has a marginal contribution to the total metal accumulation in the

CW. However, it has to be noted that metal concentrations in Phragmites australis roots were

not determined. Metal concentrations in roots are usually higher than metal concentrations in

aboveground plant parts (Vymazal, 2003). Therefore, the metal mass accumulated in the roots

is thought to be significantly higher than the mass accumulated in the aboveground plant

parts. The relative importance of the sediment could thus be overestimated in this study.

However, the relative contribution of the roots to the total metal mass accumulated in the CW

is thought to be low compared to that of the sediment. Sediments serve as sinks for metals

entering the surface flow reed. Metals appear to be efficiently removed from the wastewater

and are retained within the sediment of the CW. Lesage et al. (2005) and Vymazal (2003) also

described the high importance of metal accumulation in the sediment to the overall metal

removal in HSSF wetlands.

6.1.2. Pollution level of the sediment

The total metal concentration in the sediment gives an indication of the pollution level of the

sediment. It does however not give information on the mobility and availability of the metals.

Total metal concentrations in the sediment of the SF reed bed were presented in Table 5.9.

Important conclusions on the pollution level are summarized below:

• Total metal concentrations varied between sampling positions although no clear

trends were distinguished for most metals. Only for Zn, Ni, and Cu higher total

concentrations were seen in the upper sediment layer in the inlet area although

standard deviations were high. This study did not report a clear relationship

between the total metal concentrations in the sediment and the distance along

the treatment path.

• The total metal concentrations of the deeper sediment layer (15 – 30 cm) were

higher than those of the upper sediment layer. This was the case for all metals

except Ni.

Discussion

61

• Total metal concentrations in the sediment of the SF reed bed were elevated

compared to background values but were lower than the soil remediation

criteria (Vlarebo, 1996).

• The pollution level of the sediment is low to moderate and does not form an

immediate risk at this stage of operation.

6.1.3. Metal mobility in the sediment

Although the pollution level of the sediment is considered low to moderate, it is still

interesting to investigate the mobility of metals in the sediment. A sediment with a moderate

pollution level, but a high metal mobility can form a potential risk when sediment

characteristics change. Metal mobility is mainly affected by sediment pH, % of organic

matter, redox status, and texture as described in paragraph 2.2.2. The mobility of metals was

investigated by means of two parameters: the exchangeable concentration levels in the

sediment and the SEM/AVS ratio.

6.1.3.1. Metal mobility assessed by the exchangeable metal fraction

Single extraction procedures are used to identify different operationally defined fractions. The

exchangeable fraction consists of metals that can be readily released into the soil solution by

cation exchange processes. Exchangeable metals represent a labile metal fraction and give an

idea of the metal mobility in the sediment. A comparison of the exchangeable metal

concentrations to the total metal concentrations was performed and presented in Table 5.11.

Following order of decreasing exchangeability was noted:

Cd > Mn > Zn > Pb ~ Ni > Cu > Cr > Fe

Cd, Mn, and Zn appeared to be the most mobile metals in the sediment with 15 – 39, 12-14,

and 6-10 % of the total metal content being exchangeable respectively. These elements are

considered to be the most mobile ones. Less than 0.01 % of the total Fe present in the

sediment was present in an exchangeable state. Fe is therefore not considered to be a labile

element.

Discussion

62

6.1.3.1. Metal mobility assessed by the SEM/AVS ratio

In the strongly anoxic sediments of the SF reed bed in Deurle, the redox potential is thought

to be a major factor regulating metal mobility. The redox potential of the upper sediment layer

(0 – 15 cm) varied between -150 and -200 mV (Fig. 5.4.). A similar redox potential was seen

in the deeper sediment layer (15 – 30 cm) further in the SF wetland whereas lower redox

potentials (between -300 and -350 mV) where seen in the inlet area. Reported redox potentials

indicate that reduction of SO42- is taking place. Indeed, AVS were detected in the wetland

sediments. Similar AVS levels were seen in both sediment layers from a distance of 70 m

from the inlet onward (Fig. 5.5.). However, contradictory to what one would expect higher

AVS levels were reported in the upper sediment in the inlet area than in the deeper sediment.

An important removal process for trace metals in SF wetlands is the binding of trace metals to

wetland substrates as insoluble compounds, particularly as metal sulphides (Scholes et al.,

1998). In the reduced sediments of the SF wetland, metals can be precipitated by sulphides.

The SEM/AVS ratio can be considered as a measure for the potentially bioavailable metal

fraction (van den Hoop et al., 1997). The SEM/AVS ratio’s of both sediment layers at

different sampling positions was lower than 1 (Table 5.13.), indicating that metals are bound

by sulphides. Cd, Cu, Ni, Pb, and Zn are assumed to be immobilized in the sediment. When

Fe and Mn are included in the SEM/AVS ratio, the latter always exceeded 1. This would

indicate that Fe and Mn are potentially bioavailable.

6.1.3.3. Applicability of the SEM/AVS ratio

As the SEM/AVS ratio < 1, Cd, Cu, Ni, Pb, and Zn are assumed to be immobilized. However,

when the exchangeable metal concentrations are considered, different conclusions were

deducted. 15-39 and 6-10 % of the Cd and Zn respectively present in the sediment was

exchangeable and can be considered mobile. This can be explained by the fact that the

SEM/AVS principle assumes that all extracted metals are present as metal sulphides, an

assumption somewhat arbitrary as the reagent does not exclusively attack the metal sulphides.

When Fe and Mn are included in the SEM/AVS ratio, the ratio is > 1 at each sampling

position, indicating that Fe and Mn are potentially available. 12-14 % of the Mn present in

the sediment is exchangeable (Table 5.11.). This corresponds to the SEM/AVS ratio > 1 and

Discussion

63

the idea that Mn is potentially available. However, when Fe is considered, it is noted that less

than 0.01 % of the total Fe content is exchangeable. However, the SEM/AVS ratio would

indicate that Fe is potentially available. Fe can be bound by other fractions such as carbonates

and organic matter. Results suggest that the SEM/AVS ratio should be used as a measure of

potential availability but not as a strict measure. Van den Hoop et al. (1997) observed

seasonal variations in AVS and SEM concentrations which leads to variation in metal

availability for organisms during the year.

6.1.3.4. Metal mobility and implications for future use of the SF wetland

Mobility of metals in the sediment of the SF wetland is generally low. The SEM/AVS ratio is

lower than 1 and indicates that Cd, Cu, Ni, Pb, and Zn are not potentially available. However,

15-39, 76-10, and 12-14 % of the Cd, Zn, and Mn is exchangeable, respectively. Less than 5

% of the total Pb, Ni, and Cu level, less than 0.3 % of the total Cr level, and less than 0.01 %

of the total Fe level is exchangeable.

In the reduced sediments of the SF wetland, metals can be precipitated by sulphides. As long

as the sediment remains reduced, metals remain in an immobile state. When for example the

field would be aerated and would not be used for wastewater treatment anymore, then

possible mobilization of metals could occur. If sediments are drained and aerated, metal

sulphides will get oxidized and the solubility of metals will increase. Moreover, the pH can

decrease if the buffer capacity of the sediment is low. This could result in a large availability

of metals for soil living organisms, leaching and plant uptake (Harmsen, 2004). For the time

being, leaching of metals should not represent a problem as strongly reduced conditions in

wetland soils favor the immobilization of metals (Gambrell, 2004).

6.2. Removal efficiency of the surface flow wetland

Wetlands are capable of removing large quantities of trace metals from the wastewater (Ye et

al., 2001). There was considerable variation in the concentration of each trace metal in the

wastewater between both sampling dates (paragraph 5.1.2). Generally, concentration levels

were lower in October 2004, which might be attributed to the dilution of the wastewater

because of rainfall at the time of sampling.

Discussion

64

Although dissolved metal concentrations in the wastewater were low, one could observe a

clear decrease along the treatment path of the SF wetland for Al, Cu, and Zn at both sampling

dates. High removal efficiencies were noted in October 2004 for Al, Cu, Fe, Pb, and Zn, being

28, 64, 32, 40, and 57 % respectively. During the 2nd sampling in November 2004, high

removal efficiencies were seen for Al, Cu and Zn (63, 75, and 62 % respectively).

Dissolved Mn concentrations increased with increasing distance from the inlet, at both

sampling dates. Elevated concentrations of Mn in the wastewater may be explained by

reduction processes occurring in the wetland sediment. Mn is known to be present in surface

waters as Mn (IV) and the relatively unstable Mn (III), which forms insoluble oxides and

hydroxides. At low redox potentials and low pH, the predominant form is Mn (II) (Kadlec &

Knight, 1996). AVS levels in the sediment are high, indicating the importance of sulphides in

the immobilization of metals. However, SEM/AVS ratios with Fe and Mn included are > 1 at

each sampling position, indicating that Fe and Mn are potentially available. It is reminded that

MnS has the highest solubility product compared to other metal sulphides (Table 2.6.). When

exchangeable Mn concentrations are compared to total Mn concentrations in the sediment

(Table 5.12.) it seems that 12-14 % of the Mn present in the sediment is readily exchangeable.

Mn can thus be considered to be a mobile element. Mn appears to be released from the

sediment and migrates through the SF wetland.

The removal efficiency of Cd, Cr, and Ni was generally low, mainly because of the very low

concentration levels of these metals in the influent wastewater. Rousseau et al. (2004)

mentioned that influent concentrations of COD, BOD and nutrients in the SF constructed

wetlands are the lowest ones compared to other types of CW’s in Flanders. This could be

attributed to the combination of domestic wastewater with rainwater and could also be a

reason of the very low dissolved concentration levels of some trace metals. Vymazal (2001)

reported very high removal efficiencies for Pb (98 %), Ni (92 %), and Cd (77 %) when

influent concentrations were higher. Scholes et al. (1998) reported total removal efficiencies

of 68 % for Cu and 65 % for Pb at the Brentwood wetland compared to Dagenham, which had

negative removal efficiencies of respectively –180 % and –171 %. Both of these CW’s treat

urban runoff in the UK.

Discussion

65

It has to be emphasized that the constructed wetland is one of the oldest constructed wetlands

in Flanders, being in operation for 16 years. Monitoring data that are reported in this research

cover only two sampling times in a two-month period, which is not enough to make

straightforward conclusions regarding the removal efficiency of the CW. Therefore, regular

monitoring of the influent wastewater and the effluent is needed in order to assess the removal

efficiency of the CW.

Conclusions

66

7. CONCLUSIONS This thesis aimed at assessing heavy metal accumulation in a SF reed bed that has been in

operation for 16 years. Metal concentrations in the wastewater, sediment, and aboveground

Phragmites australis biomass were investigated. Exchangeable metal concentrations and

SEM/AVS ratio were determined in order to assess metal mobility and potential availability.

The relative contribution of the aboveground Phragmites australis biomass to the overall

metal accumulation in the CW was of marginal importance. Metals were mainly accumulated

in the sediment of the SF wetland. Total metal concentrations in the sediment were not clearly

affected by sampling position along the treatment path, although higher Zn, Ni, and Cu

concentrations could be detected in the inlet area of the bed. Total metal concentrations in the

deeper sediment layer (15 – 30 cm) were generally higher than in the top sediment layer (0-15

cm). The pollution level was low to moderate and soil remediation criteria were not exceeded.

The pollution level itself does not give sufficient information on the mobility of the metals

and the potential risk. The SEM/AVS ratio (Cu+Pb+Cd+Ni+Zn) is a measure of the potential

availability and was lower than 1 at each sampling position. This indicates that heavy metals

were present in a non-available form and that they were precipitated as metal sulphides in the

sediment. A SEM/AVS ratio < 1 means that there may be no acute toxicity for aquatic

organisms in terms of these heavy metals. On the contrary, SEM/AVS was > 1 when Fe and

Mn were included, what might indicate that Fe and Mn are potentially mobile. However, use

of the SEM/AVS principle on its own without the assessment of additional mobility

parameters is questioned. The exchangeable metal fraction gives additional information

towards mobility. Cd, Zn, and Mn appear to be mobile as 15-39, 6-10, and 12-14 % of the

total metal content was exchangeable, respectively. Less than 5 % of the total Pb, Ni, and Cu

level, less than 0.3 % of the total Cr level, and less than 0.01 % of the total Fe level was

exchangeable, indicating a low mobility of these elements.

At this stage of operation there is no problem with metal accumulation in the sediment of the

SF wetland in Deurle, as the metal mobility in the sediment is generally low and the pollution

level is low to moderate. However, accumulation continues to take place with the operational

lifetime of the reed bed and metal mobility can be affected by changing sediment

characteristics. As long as the sediment remains reduced, metals remain in an immobile state.

Conclusions

67

When the field would be aerated or if sediments would be dredged and aerated, metal

sulphides will get oxidized and the solubility of metals will increase. This change in land use

may pose a threat to the environment as once oxidized, sediment pH could drop, acidifying

conditions could be created, and mobility of metals would rise. Therefore, future applications

of the reed bed must be adequately considered and specified before any changes in use or

operation are performed.

Metal concentrations in aboveground plant parts of Phragmites australis were not affected by

sampling position in the CW. Highest concentrations of Al, Cu, Ni, Pb, and Zn were seen in

the panicles, whereas highest concentrations of Fe and Mn were seen in the leaves. The metal

mass accumulated in the aboveground biomass has a marginal contribution to the total mass

accumulated in the CW. Future research could include the analysis of metals in the roots of

Phragmites australis.

Although dissolved metal concentrations in the wastewater were low, one could observe a

clear decrease along the treatment path of the SF wetland for Al, Cu, and Zn at both sampling

dates. The reed bed showed to be highly efficient in the removal of Al, Cu, and Zn, indicating

28, 63, and 57 % in October 2004 and 63, 78, and 62 % in November 2004, respectively. The

removal efficiency of Cd, Cr, and Ni was generally low, mainly because of the very low

concentration levels of these metals in the influent wastewater. Dissolved Mn concentrations

increased with increasing distance from the inlet, at both sampling dates. This was attributed

to reducing conditions in the sediment. More regular monitoring of the influent wastewater

and the effluent is needed in order to obtain a representative picture of the removal efficiency

of the surface flow wetland.

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75

APPENDICES

APPENDIX I Table a – i: Metal concentration in different plant parts as a function of

distance from the inlet expressed in mg kg-1

Table a. Al concentration in Phragmites australis plant parts (mg kg-1)

Position (m) Stems Leaves Leaf sheats Panicles 0 8 ± 1.31 15 ± 2.16 12 ± 1.09 28 ± 0.91

25 7 ± 2.18 20 ± 1.75 12 ± 0.93 20 ± 3.10 70 11 ± 5.61 14 ± 0.50 14 ± 0.51 18 ± 1.97 140 7 ± 0.60 16 ± 1.24 11 ± 0.25 18 ± 0.85 350 9 ± 0.46 25 ± 0.60 17 ± 1.02 87 ± 12.41 630 8 ± 1.22 16 ± 1.36 11 ± 1.55 31 ± 12.82

Table b. Cd concentration in Phragmites australis plant parts (mg kg-1)

Position (m) Stems Leaves Leaf sheats Panicles 0 0.03 ± 0.01 0.07 ± 0.00 0.07 ± 0.00 0.06 ± 0.03

25 0.02 ± 0.01 0.06 ± 0.01 0.04 ± 0.03 0.05 ± 0.03 70 0.04 ± 0.01 0.05 ± 0.02 0.05 ± 0.02 0.04 ± 0.03

140 0.05 ± 0.01 0.04 ± 0.02 0.05 ± 0.01 0.04 ± 0.02 350 0.03 ± 0.04 0.05 ± 0.01 0.05 ± 0.02 0.06 ± 0.02 630 0.04 ± 0.02 0.06 ± 0.02 0.03 ± 0.02 0.05 ± 0.02

Table c. Cr concentration in Phragmites australis plant parts (mg kg-1)


25 0.81 ± 0.06 0.75 ± 0.04 0.69 ± 0.07 0.88 ± 0.08 70 1.75 ± 0.24 0.69 ± 0.06 0.73 ± 0.02 0.76 ± 0.01 140 1.41 ± 0.07 1.08 ± 0.02 0.80 ± 0.06 0.79 ± 0.10 350 1.70 ± 0.01 0.84 ± 0.04 1.29 ± 0.16 1.53 ± 0.07 630 1.23 ± 0.09 0.71 ± 0.03 0.63 ± 0.05 0.90 ± 0.19

Table d. Cu concentration in Phragmites australis plant parts (mg kg-1)


25 0.72 ± 0.10 2.26 ± 0.04 1.21 ± 0.03 2.94 ± 0.72 70 1.08 ± 0.08 3.47 ± 0.09 1.82 ± 0.05 5.52 ± 0.55 140 1.57 ± 0.12 3.60 ± 0.59 1.92 ± 0.11 4.52 ± 0.39 350 1.40 ± 0.05 3.75 ± 0.07 2.18 ± 0.26 8.73 ± 0.71 630 1.33 ± 0.03 5.14 ± 0.06 2.02 ± 0.11 4.21 ± 0.40

76

Table e. Fe concentration in Phragmites australis plant parts (mg kg-1)


25 24 ± 5.00 179 ± 3.32 53 ± 1.67 111 ± 10.19 70 37 ± 4.61 127 ± 2.19 59 ± 0.47 102 ± 1.51 140 32 ± 1.82 146 ± 8.96 47 ± 2.16 103 ± 1.34 350 46 ± 1.95 158 ± 4.38 73 ± 1.43 297 ± 35.95 630 30 ± 1.96 158 ± 1.09 52 ± 4.67 130 ± 32.95

Table f. Mn concentration in Phragmites australis plant parts (mg kg-1)


25 14 ± 1.62 159 ± 1.84 58 ± 0.70 64 ± 13.84 70 30 ± 1.17 296 ± 4.33 119 ± 4.17 76 ± 11.88 140 45 ± 1.25 177 ± 10.35 99 ± 6.47 81 ± 1.34 350 41 ± 1.20 263 ± 3.21 115 ± 1.36 74 ± 2.08 630 16 ± 1.35 124 ± 0.80 49 ± 1.23 81 ± 9.63

Table g. Ni concentration in Phragmites australis plant parts (mg kg-1)


25 0.27 ± 0.12 0.67 ± 0.08 0.52 ± 0.13 0.54 ± 0.02 70 0.58 ± 0.09 0.62 ± 0.10 0.88 ± 0.12 1.07 ± 0.21 140 0.59 ± 0.05 0.95 ± 0.01 0.82 ± 0.03 0.70 ± 0.26 350 0.73 ± 0.16 0.75 ± 0.04 0.90 ± 0.10 1.06 ± 0.09 630 0.43 ± - 0.44 ± 0.03 0.65 ± 0.07 0.66 ± 0.08

Table h. Pb concentration in Phragmites australis plant parts (mg kg-1)


25 0.69 ± 0.30 0.81 ± 0.35 0.46 ± 0.16 0.92 ± 0.09 70 0.59 ± 0.13 1.23 ± 0.10 0.88 ± 0.27 1.41 ± 0.05 140 0.62 ± 0.07 1.24 ± 0.26 0.65 ± 0.54 1.46 ± 0.04 350 0.27 ± 0.40 1.35 ± 0.27 0.72 ± 0.16 3.80 ± 0.87 630 0.26 ± 0.24 1.05 ± 0.50 0.99 ± 0.32 1.65 ± 0.74

Table i. Zn concentration in Phragmites australis plant parts (mg kg-1)


25 41.64 ± 9.15 40.85 ± 1.07 39.73 ± 0.21 70.05 ± 4.25 70 112.08 ± 6.51 48.17 ± 0.72 44.54 ± 2.99 104.22 ± 0.81 140 53.91 ± 1.69 57.32 ± 4.70 37.72 ± 1.88 84.72 ± 7.69 350 98.83 ± 1.65 50.74 ± 0.89 45.47 ± 3.87 87.71 ± 4.47 630 41.20 ± 3.34 41.62 ± 0.37 33.57 ± 1.51 60.85 ± 5.66

77

APPENDIX II. Figure. a – h: Total metal concentration in both sediment layers of the SF

reed bed as a function of distance from the inlet expressed in mg kg-1

Figure a. Total Cr concentration in the sediment

020406080

100120140160180

0 25 70 140 350 630Distance (m)

Cr m

g kg

-1

Cr top layer

Cr bottom layer

Background value

Soil sanitation standard

Figure b. Total Cu concentration in the sediment

0

50

100

150

200

250

300

0 25 70 140 350 630Distance (m)

Cu

mg

kg-1

Cu top layer

Cu bottom layer

Background value

soil sanitation standard

Figure c. Total Cd concentration in the sediment

0

1

2

3

4

5

6

7

8

0 25 70 140 350 630

Distance (m)

Cd

mg

kg-1

Cd top layerCd bottom layerBackground valueSoil sanitation standard

78

Figure d. Total Fe concentration in the sediment

0

5000

10000

15000

20000

25000

0 25 70 140 350 630

Distance (m)

Fe m

g kg

-1Fe top layer

Fe bottom layer

Figure e. Total Mn concentration in the sediment

0

50

100

150

200

250

300

0 25 70 140 350 630

Distance (m)

Mn

mg

kg-1

Mn top layer

Mn bottom layer

Figure f. Total Ni concentration in the sediment

0

20

40

60

80

100

120

140

0 25 70 140 350 630

Distance (m)

Ni m

g kg

-1

Ni top layer

Ni bottom layer

Background values


b

79

Figure g. Total Pb concentration in the sediment

0

100

200

300

400

500

600

700

800

0 25 70 140 350 630Distance (m)

Pb m

g kg

-1Pb top layer

Pb bottom layer

Background value

soil sanitation standard

Figure h. Total Zn concentration in the sediment

0

200

400

600

800

1000

1200

0 25 70 140 350 630Distance (m)

Zn m

g kg

-1

Zn top layer

Zn bottom layer

Background values


80

APPENDIX III. Table a – i: Simultaneously extracted metals (SEM) in both sediment layers

of the SF reed bed as a function of distance from the inlet expressed in mg kg-1

Table a. SEM - Al concentration (mg kg-1)

Position (m) 0-15 cm 15-30 cm 0 3551 ± 2619 724 ± 143

25 1323 ± 525 1165 ± 947 70 875 ± 96 1404 ± 977

140 1068 ± 370 2115 ± 652 350 934 ± 147 1271 ± 197 630 745 ± 106 1029 ± 237

Table b. SEM - Cd concentration (mg kg-1)

Position (m) 0-15 cm 15-30 cm 0 1.87 ± 0.99 0.70 ± 0.08

25 1.48 ± 0.64 1.40 ± 1.60 70 0.98 ± 0.33 3.64 ± 3.12

140 1.12 ± 0.62 2.51 ± 0.62 350 0.63 ± 0.34 1.00 ± 0.58 630 1.04 ± 0.17 1.57 ± 0.56

Table c. SEM – Cr concentration (mg kg-1)

Position (m) 0-15 cm 15-30 cm 0 16 ± 6 6 ± 1

25 14 ± 8 13 ± 13 70 12 ± 4 33 ± 26

140 15 ± 10 34 ± 12 350 14 ± 2 24 ± 4 630 8 ± 2 14 ± 6

Table d. SEM - Cu concentration (mg kg-1)

Position (m) 0-15 cm 15-30 cm 0 - ± - 0.6 ± -

25 0.64 ± - 1.3 ± 1 70 1.49 ± 0.5 18 ± 19 140 - ± - 59 ± 30 350 0.07 ± - 1.2 ± 1.5 630 15.50 ± 7.4 26 ± 25

81

Table e. SEM – Fe concentration (mg kg-1)

Position (m) 0-15 cm 15-30 cm 0 6899 ± 2805 3423 ± 412

25 5551 ± 1876 4105 ± 3123 70 2333 ± 503 4430 ± 2959

140 3552 ± 1265 8355 ± 2548

350 4492 ± 459 5978 ± 1867 630 2067 ± 491 3046 ± 787

Table f. SEM - Mn concentration (mg kg-1)

Position (m) 0-15 cm 15-30 cm 0 123 ± 34 68 ± 28

25 114 ± 34 106 ± 77 70 61 ± 11 125 ± 89

140 97 ± 14 200 ± 15 350 104 ± 20 158 ± 97 630 75 ± 26 116 ± 14

Table g. SEM - Ni concentration (mg kg-1)

Position (m) 0-15 cm 15-30 cm 0 13 ± 7 5 ± 1

25 8 ± 4 7 ± 7 70 6 ± 1 18 ± 18

140 7 ± 4 12 ± 3 350 8 ± 1 13 ± 3 630 7 ± 3 5 ± 1

Table h. SEM - Pb concentration (mg kg-1)

Position (m) 0-15 cm 15-30 cm 0 121 ± 75 51 ± 21

25 102 ± 53 88 ± 91 70 57 ± 25 184 ± 161

140 127 ± 82 252 ± 97 350 73 ± 15 81 ± 11 630 61 ± 19 97 ± 42

Table i. SEM - Zn concentration (mg kg-1)

Position (m) 0-15 cm 15-30 cm 0 610 ± 588 133 ± 20

25 252 ± 84 234 ± 233 70 172 ± 68 580 ± 466

140 216 ± 91 565 ± 155 350 208 ± 50 265 ± 83 630 162 ± 72 562 ± 694

82

GLOSSARY

Adsorption The adherence of a gas, liquid or dissolve chemical to the

surface of a solid, eg. sediment particle

Aggregation Process whereby small particles cluster together due to

particle attraction forces

Aerobic A state where free oxygen is available

Anaerobic A state where neither free oxygen nor oxygen bound to other

molecules is available

Anoxic A state where there in no free oxygen, but oxygen bound to

other molecules is available

Aspect ratio The ratio of the wetland length to its width

Benthic Occurs on or in the bottom sediments of a wetland

Biofilm An organic layer, typically composed of algae, micro fauna

and bacteria, which adsorb small particles (colloids) and

nutrients. Biofilms are important treatment component

within CW’s

Biomass The living weight of plants or animals

Biodiversity See Diversity

BOD Measurement of the oxygen consumed during bacterial

breakdown of organic matter in water

Constructed wetland A wetland with a purpose to achieve certain treatment using

soil, water and biota

Desorption The release back into solution of substances that have been

previously adsorbed onto a surface

Detention time The average period of time that effluent is detained within

the wetland

Detritus Dead plant material that is in the process of microbial

decomposition

Diversity The number and distribution of animal and plant species

within a defined area

Effluent A liquid that flows out of a process or treatment system

83

Eh A measure of redox potential (oxidation-reduction potential)

expressed in mV

Emergent plants Plants that are attached to the substrate and whose leaves

and stems either float or protrude above the surface

Hydraulic conductivity Rate at which soil or substrate can transmit water

Hydraulic loading rate Influent discharge into wetland per square meter of wetland

surface

Hydraulic residence time See Detention time

Infiltration The process of water moving into the surface of the soil or

substrate

Influent A liquid that flows into a process or treatment system

Macrophyte Plants that are macroscopic, i.e. visible to the naked eye

Maturation pond A pond used to treat secondary effluent. These ponds

generally receive low effluent loads, are aerobic and have

long retention times

Nitrification Biological process by which bacteria convert ammonia to

nitrate nitrogen

Oxidation The addition of oxygen to a substance, or the removal of

hydrogen from it. Reaction in which an atom losses an

electron

Oxidation pond Or stabilization pond is a general term for various pond

systems used in wastewater treatment. These ponds could be

aerobic, anaerobic or include both aerobic and anaerobic

conditions

pH A measure of the hydrogen ions concentration in a solution,

indicating the presence of acidic, neutral or alkaline

conditions

Precipitation Chemical reaction causing substance in a solution to be

deposited as a solid

Redox The potential of sediments to oxidize or reduce chemical

substance. A redox potential Eh>300 mV indicates aerobic

conditions, and Eh<-100 mV indicates anaerobic conditions

Reduction Reaction in which an atom accepts an electron

84

Reuse The beneficial use of treated wastewater

Retention time See Detention time

Rhizosphere The chemical sphere of influence of plant roots in soils

Substrate

Material that forms the wetland bed and provides the base

for wetland planting

Surface flow wetland Wetland designed to have water surface above the wetland

bed or substrate. Also referred as Free surface, free water

surface, open water surface wetlands

Sub-surface flow wetland Wetland designed that the flow moves through the soil or

gravel matrix, which is planted with macrophytes

Submerged plants Plants that may be attached to the wetland substrate or free

floating, but whose leaves and stems are permanently

submerged under water

Volatilization Conversion of a chemical substance from a liquid or solid to

a gas

Water balance Water volume changes in a wetland in response to variations

in wastewater discharges, rainfall, seepage,

evapotranspiration and other hydrological factors

85

Maja Šimpraga was born in Zagreb (Croatia) on 23 September 1977. She attended Bukovac

Elementary School, and III Gimnazija High School in Zagreb. In 1996 she graduated form

Oroville High School (Oroville, CA, USA). In 2003 she graduated from the University of

Zagreb, Faculty of Agriculture with a thesis title ‘Economically important pest - cherry fruit

fly, Rhagoletis cerasi L.’, obtaining a degree of agricultural engineer in phytomedicine. In

2002 she received rector’s award on voluntary research ‘Problem of the pear rust,

Gymnosporangium sabinae DC. in the urban areas’. In April 2002 she attended a monthly

training at the Royal Research Institute of Gorsem (‘Koninklijk Opzoekingsstation van

Gorsem’), Department of Entomology and Mycology, in St. Truiden (Belgium). She was an

active member of International Association of Agriculture Students. Since September 2003

she attends University of Ghent, Ghent (Belgium), majoring in Environmental Sanitation.

metal accumulation in a constructed surface flow...

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