Faculty of Bioscience Engineering

Academic year 2013 – 2014

Fate of silver nanoparticles in treatment wetlands

Viviana Vásquez Sepúlveda

Promotor: Prof. ir. Diederik Rousseau

Prof. ir. Gijs Du Laing

Tutor: ir. Reeta Auvinen

Master’s dissertation submitted in partial fulfillment of the requirements for

the degree of

Master of Science in Environmental Sanitation

i

COPYRIGHT

The author, the promoter and the tutor give permission to use this thesis for consultation and to

copy parts of it for personal use. Any other use is subject to the Laws of Copyright. Permission to

produce any material contained in this work should be obtained from the author.

© Gent University, June 2014

The Promoters:

Prof. ir. Gijs Du Laing

Prof. Ir. Diederik Rousseau

The Tutor:

ir. Reeta Auvinen

The Author:

Viviana Vásquez Sepúlveda

ii

Acknowledgment

I would like to show my gratitude to my promoter, Prof. Gijs Du Laing and Prof. Diederik Rousseau for

giving me the opportunity to work in the Laboratory of Analytical Chemistry and Applied

Ecochemistry (Ecochem) as a master’s student, and also for his valuable comments and suggestions

during the whole thesis work.

I would like to give my sincere thanks to my tutor, Hannele Auvinen for her excellent supervision,

friendly assistance and valuable recommendations towards my thesis work. She was always available

for my questions and has lead me to the result presented here, the successful finalization of my

thesis work.

Special thanks to my parents and sister, for their love, trust and all their support in my life, for letting

me spread my wings and fulfill my dreams. To my friends back home for all their support and

company, even at a distance. To my friends here in Gent, for being their constant support and

company during these two years, for making this journey unforgettable.

Lastly, to CONICYT and their Becas Chile program and the board of the Environmental Sanitation

Master program for selecting me and giving me the opportunity to study here.

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Abstract

Silver nanoparticles are commonly used in consumer products, such as personal care products, paint,

fabrics and disinfectants. Although the toxicity of silver has been well studied, the toxicity of silver

nanoparticles in the environment is still not well understood, due to their special properties

compared to the bulk material. Silver nanoparticles contained in consumer products are mainly

released to wastewater systems; hence it is important to understand the fate of these particles in

wastewater treatment facilities to avoid unintentional releases of silver nanoparticles to the natural

environment. Constructed wetlands are a good alternative to conventional wastewater treatment

plants, offering a series of advantages. Several processes act in the removal of pollutants in

constructed wetlands, such as adsorption, chemical reactions, uptake by biofilm and plants and

filtration. In this thesis, different experiments were performed in order to evaluate the removal of

silver nanoparticles in constructed wetlands and each process was assessed individually. Adsorption

experiments were performed in solid wetland substrates: sand, gravel and zeolite, and the effect of

biofilm formation and activity were also analyzed measuring the adsorption and uptake during 24

hours. To assess the adsorption and uptake of wetland plants (P. australis) an experiment was

performed during 4 weeks and a mass balance of the total amount of silver added in that period was

done. The adsorption on solid substrates was proven to be higher on sand and zeolites compared to

gravel, both at 100 µg Ag-NPs/l and 1000 µg Ag-NPs/l of initial concentration. Significant differences

in the amount of adsorbed silver were found between sand and gravel and zeolite and gravel. All in

all, the effect of adsorption onto the substrates was found low compared to the substrate sample

containing biofilm (0.21±0.03 µg Ag/g substrate versus 0.06±0.01 µg Ag/g substrate). The presence of

an active biofilm on gravel enhanced the removal of silver nanoparticles from wastewater by

adsorption and uptake processes by 350%. Significant differences were found between adsorption on

active biofilm and no biofilm; however the passivation of biofilm by drying at 40°C shows a decrease

in adsorption and no significant differences compared with adsorption in substrates without biofilm.

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This suggests that the uptake of silver nanoparticles by microorganisms is the predominant process,

or that the dry biofilm was detached from the gravel surface in contact with water. In the experiment

conducted with plants it was found that the wetland plants were able to adsorb/uptake silver

nanoparticles from the water phase to some extent. However after the third week an increase in the

concentration of silver in the water phase was found, suggesting that the maximum adsorption

capacity of the plant was reached or that desorption of Ag from the roots took place. The uptake of

silver by roots also contributes to the removal of silver nanoparticles from the water phase, but this

amount is smaller or equal to the adsorption process (13% versus 19% of the total mass of silver

added in the system in passive plants, and 19% versus 19% of the total mass in the case of active

plants). In short, constructed wetland components were proven to act as a sink for silver

nanoparticles during wastewater treatment, via processes of adsorption and uptake by solid

substrates, biofilm and plant roots. Although the present thesis gives an insight in the fate of silver

nanoparticles in constructed wetlands for wastewater treatment, it is necessary to further study

these processes combined and during a longer period of time.

Keywords: constructed wetlands, silver, adsorption, biofilm, plant uptake

v

Table of contents

Contents Page

Copyright i

Acknowledgment ii

Abstract iii

Table of contents v

List of tables vii

List of figures viii

List of abbreviations ix

1. Introduction 1

1.1 Background 1

1.2 Objective of the study 2

2. Literature Review 3

2.1 Nanoparticles 3

2.2 Environmental occurrence and effects of silver nanoparticles 5

2.3 Removal of Ag-NPs during wastewater treatment 8

2.3.1 Removal Processes 8

2.3.1.1 Adsorption mechanisms 8

2.3.1.2 Sulfidation and precipitation 9

2.3.1.3 Aggregation of Ag-NPs 11

2.3.1.4 Bioaccumulation by microorganisms and plant uptake 12

2.3.2 Conventional wastewater treatment plants 14

2.3.3 Constructed wetlands for wastewater treatment 17

3. Materials and Methods 24

3.1 Synthetic wastewater 24

3.2 Characterization of nanoparticles and substrates 25

3.2.1 Characterization of Ag-NPs solution 25

3.2.2 Characterization of substrates 25

3.2.2.1 pH 25

3.2.2.2 Electrical conductivity 25

3.2.2.3 Organic matter content 26

vi

3.2.2.4 Cation exchange capacity (CEC) 26

3.2.2.5 Determination of trace metals 26

3.3 Adsorption experiments 27

3.3.1 Preliminary test to evaluate the effect of centrifugation speed 28

3.3.2 Preliminary test to evaluate the effect of wastewater strength 28

3.3.3 Adsorption on solid substrates 29

3.3.4 Effect of biofilm formation 29

3.4 Adsorption to plant material and uptake by wetland plants 30

3.5 Statistical analysis 32

4. Results 33

4.1 Characterization of nanoparticles and substrates 33

4.1.1 Characterization of Ag-NPs solution 33

4.1.2 Characterization of substrates 33

4.2 Effect of centrifugation speed and wastewater strength 35

4.3 Adsorption of Ag-NPs to different wetland substrates 36

4.4 Influence of biofilm on removal of Ag-NPs from wastewater 37

4.5 Adsorption to plant material and uptake by the plants 39

5. Discussion 42

5.1 Effect of centrifugation speed and wastewater strength 42

5.2 Adsorption of Ag-NPs to different wetland substrates 43

5.3 Influence of biofilm on adsorption / uptake by micro-organisms 45

5.4 Plant adsorption and uptake 48

5.5 General assessment of the removal processes in constructed wetlands 50

6. Conclusions and Recommendations 52

References 54

vii

List of Tables

Page

Table 2.1. Solubility products (Ksp) of silver containing solids (adapted from

Levard & Hotze 2012).

11

Table 2.2. Estimated distribution of mass of silver recovered in mesocosms

dosed in either the terrestrial compartment or the water column

(Lowry, Espinasse, et al. 2012).

23

Table 3.1 Concentration of the constituents of synthetic wastewater used in the

experiments.

24

Table 4.1. Main physicochemical characteristics of different wetland substrates

(average ± SD, n=3).

34

Table 4.2. Trace metals concentrations in sand, gravel and zeolite determined via

ICP-OES after aqua regia digestion (average ± SD, n=3).

34

Table 4.3. Mass balance of Ag-NPs during a 4-week experiment focused on

adsorption to wetland plants and uptake by the plants. Amounts of

silver in- and out-of the system, considering all the phases involved

(average ± SD, n=3).

40

viii

List of Figures

Page

Figure 2.1 A typical wastewater treatment plant including preliminary

treatment, primary treatment, secondary treatment, tertiary

treatment and solids treatment and disposal (Brar et al. 2010).

16

Figure 2.2 Classification of constructed wetlands for wastewater treatment

and scheme of the operation of free floating plants, surface flow

and sub-surface - horizontal and vertical - flow constructed

wetlands (top to bottom) (Vymazal 2007).

20

Figure 3.1 Scheme of adsorption experiments performed with solid

substrates, and subsequent sample preparation and analysis.

27

Figure 3.2 General scheme of the different compounds taken into account to

construct the mass balance of Ag.

32

Figure 4.1 Size distribution of silver nanoparticles solution analyzed by

Photon Correlation Spectrometry. Left: Size distribution based on

intensity; Right: Size distribution based on number.

33

Figure 4.2

Effect of the centrifugation speed on the concentration of Ag in the

supernatant after 24 hours of mixing followed by 10 minutes of

centrifugation.

35

Figure 4.3 Effect of the strength of synthetic wastewater in solution on the

concentration of Ag in the water phase after a 24 hours mixing

experiment.

36

Figure 4.4 Amounts of Ag-NPs adsorbed to sand, gravel and zeolite after a 24

hours, at different initial concentrations of Ag-NPs..

37

Figure 4.5 Effect of biofilm condition on silver adsorbed per gram of

substrate.

38

Figure 4.6 Effect of biofilm condition present in substrate on silver adsorbed

per gram of biofilm.

39

Figure 4.7 Distribution of Ag-NPs between different phases after exposing the

NPs for 4 weeks to wetland plants; top: Active wetland plant;

Bottom: Passive wetland plant.

41

ix

List of Abbreviations

Ag-NPs Silver nanoparticles

BOD Biochemical oxygen demand

CEC Cation exchange capacity

CW Constructed wetlands

CWTS Constructed Wetland Treatment System

DOC Dissolved organic carbon

DOM Dissolved organic matter

DO Dissolved oxygen

EC Electrical conductivity

EPS Extracellular polymeric substances

ICP-MS Inductively coupled plasma-mass spectrometry

ICP-OES Inductively coupled plasma-optical emission spectrometry

NP Nanoparticles

NOM Natural organic matter

PCS Photon Correlation Spectrometer

PVP Polyvinylpyrrolidone

SD Standard deviation

SSF Sub- surface flow

SF Surface flow

TSS Total suspended solids

WTR Water treatment residuals

WWT Wastewater treatment

WWTP Wastewater treatment plant

1

1. Introduction

1.1 Background

Nanoparticles (NP) are materials with at least one dimension smaller than 100 nm. This characteristic

that gives them special properties compared to the bulk material. Due to their properties, the

production and subsequent release in wastewater of nanoparticles, including silver nanoparticles

(Ag-NPs), has increase (Ju-Nam & Lead 2008; Christian et al. 2008). Silver nanoparticles are one of the

most used in consumer products, especially due to their bactericidal properties. Currently, more than

400 products contain Ag-NPs, such as paint, fabrics, disinfectants and personal care products (Sheng

& Liu 2011). The release of Ag-NPs from consumer products will most likely end up in wastewater

treatment (WWT) system. For this reason it is important to understand the different processes

involved in the removal of Ag-NPs in WWT systems and avoid that Ag-NPs reach the environment

(Brar et al. 2010).

Constructed wetlands (CW) are alternative WWT systems, with numerous advantages over

conventional system (Davis n.d.; Wetlands International - Malaysia Office 2003). In CW the removal

of contaminants can be achieved by different processes such as adsorption to solid substrate, organic

matter and plant roots, uptake by biofilm and plants present in the system and physicochemical

processes as precipitation, coagulation and filtration. Even though the removal of Ag-NPs in

conventional WWT plants has already received attention in the literature, other WWT systems, such

as CW, are less studied. Since the toxic effects that Ag-NPs may exert in the natural environment are

still uncertain it is of utmost importance to recognize the removal processes and the options

available to decrease their release to the natural environment.

2

1.2 Objective of the study

The aim of this study is to understand the fate of Ag-NPs in constructed wetlands for water

treatment, including the different processes that affect the removal of Ag-NPs in wetlands. For the

evaluation of the different removal processes a series of specific objectives were constructed:

1. Evaluate the adsorption of Ag-NPs to three different wetland substrates: sand, gravel and zeolite;

at different initial concentrations of Ag-NPs.

2. Asses the adsorption and/or uptake of Ag-NPs in biofilm surfaces formed in wetland substrates,

and the effect of passivation of the microorganisms.

3. Evaluate the adsorption and uptake of Ag-NPs by wetland plants and identify the most important

processes in a wastewater-plant system by performing a mass balance.

3

2. Literature Review

2.1 Nanoparticles

Nanoscience is defined as the study of materials on nanoscale level between 1 and 100 nm, and

involves the study of the formation of 2- and 3-dimensional assemblies and nanostructures or

nanomaterials (Ju-Nam & Lead 2008). A nanoparticle (NP) is a particle in which at least one of the

dimensions does not exceed 100 nm (Bystrzejewska-Piotrowska et al. 2009). Nanoparticles have a

high surface area to volume ratio, which is one of the reasons for their unusual properties (Christian

et al. 2008). This also gives special importance to the chemistry of the surface that usually is different

from the core material. In general, NP can be categorized into carbon-based particles, oxides, metals

and quantum dots (Ju-Nam & Lead 2008). Another classification for nanotechnology products

distinguishes between nanomaterials fixed on a substrate and free nanoparticles (Bystrzejewska-

Piotrowska et al. 2009).

The composition of the surface of NP is directly related to the application for which it is designed

(Christian et al. 2008). Many NP lose their properties once they have aggregated or precipitated from

the suspension. For this reason great efforts are made in preparing NP suspended in the medium of

choice, usually through the application of a surface coating that facilitates dispersion. These coatings

will influence the behavior and transformations that NP undergo in the environment (Levard & Hotze

2012). Depending on the environmental conditions, coatings will affect surface charge, aggregation

and toxicity. Different methods can be used in order to prepare stable suspensions, for instance using

small molecules that bind through covalent bonds and contain charged groups, such as citrate and

thiopropanoic acid; use of surfactants like dodecilsulfate (SDS); and use of monofunctional long chain

molecules such as amines, phosphines, carboxylates and thiols, which bind to specific sites on

nanoparticles and have long chains that impart stability (Christian et al. 2008).

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Nanotechnology is applied in agricultural, environmental and industrial sectors, in areas such as

electronics, biomedicine, pharmaceuticals, cosmetics and material sciences (Ju-Nam & Lead 2008;

Bystrzejewska-Piotrowska et al. 2009). The wide range of applications is related to the ease of

preparing and manipulating, high reactivity, large surface area, and their optical properties. It is

estimated that more than 1800 products based on nanotechnologies are already in use and that

many new products are expected to appear in the next years (Woodrow Wilson Database).

Metal oxide NPs are used in various applications related to material, chemical and biological

sciences. For example, cosmetics and sunscreens use TiO2, Fe2O3 and ZnO and fillers in dental fillings

use SiO2 (Ju-Nam & Lead 2008). Metallic NPs, such as silver and gold nanoparticles, have potential in

areas like sensing, catalysis, transport and applications in biological and medical sciences. Metallic

NPs have interesting properties such as their stability, inertness, and electrical, optical and magnetic

characteristics. Also, they have the advantage that they are easy to prepare and the possibility of

chemical modification of the surface of the nanoparticle by the use of capping agents.

The environmental uses of NPs include the treatment of surface water, groundwater and wastewater

contaminated with toxic metal ions, organic and inorganic solutes and microorganisms (Theron et al.

2008). Nanomaterials have been extensively used for rapid or cost-effective clean-up of wastes

compared to conventional approaches (Brar et al. 2010). Nanotechnology possesses the potential to

improve environmental quality, through direct application of nanomaterials to detect, prevent and

remove pollutants; as well as indirectly in the design of cleaner industrial solutions. The benefits of

using NPs are related with the physical characteristics of these: enhanced reactivity, surface area,

sub-surface transport and/or sequestration characteristics. Despite the growing application of

nanomaterials in different fields, some serious concerns have been raised in recent years within

different environmental compartments and human health (Brar et al. 2010). Effects on the

environment and human health of metals are already quite well-know, however hazards of metals

occurring in the form of nanoparticles still have to be studied (Brar et al. 2010).

5

Silver nanoparticles are one of the most used nanoparticles in consumer products, because of their

antibacterial properties. Of the 1854 consumer products that contain NPs listed in the Woodrow

Wilson Database of 2014, 410 of them contained Ag-NPs (Woodrow Wilson Database). Silver

coatings have been widely used to treat infected wounds and the prevention of biofilm formation on

home appliances (Sheng & Liu 2011). Ag-NPs have been introduced as microbial agent in fabrics and

paintings, besides they are used as disinfectants in medical institutions and research is being carried

out for their possible use in food industry and wastewater treatment. However it is their

antimicrobial properties that made them intrinsically toxic to the environment. Other nanoparticles

that present these properties are copper and titanium oxides nanoparticles (Ju-Nam & Lead 2008;

Bystrzejewska-Piotrowska et al. 2009; Fabrega et al. 2011).

2.2 Environmental occurrence and effects of silver nanoparticles

Nanoparticles can be released to the environment in different states: bare NPs, functionalized NPs,

aggregates, or embedded in a matrix (Bystrzejewska-Piotrowska et al. 2009; Gottschalk & Nowack

2011). The release can be deliberate or accidental and reach the different compartments of the

environment, i.e. water, soil and air. Subsequently they can bioaccumulate in the food chain, in case

they have a very low solubility and degradability; or undergo degradation, in which case heavy

metals may be released.

In order to predict the release and accumulation of Ag-NPs in the environment it is necessary to

know their discharge routes to the environment. This includes their discharge during synthesis,

manufacturing and incorporation of Ag-NPs into goods, use and recycling or disposal of products that

contain these compounds (Fabrega et al. 2011; Benn & Westerhoff 2008; Benn et al. 2010; Reinsch et

al. 2012; Gottschalk & Nowack 2011). Diffuse sources, resulting from use and disposal of

nanomaterial containing goods, are the most important and difficult to control. On the other hand,

point sources include production and products manufacturing facilities, transport processes,

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recycling and disposal practices from these facilities. Nanoparticles contained in liquid phases, such

as cleaning products, creams and cosmetics, are easily released during the use phase within hours,

whereas NPs embedded into a solid matrix, like coatings in fabrics and paints, are gradually released

across the whole life cycle and are supposed to remain in the matrix for years.

In general, nanomaterials used in household and industrial products are released and find their way

through waste disposal routes into WWT facilities and end up in wastewater sludge (Brar et al. 2010).

The impact that NPs have in the wastewater treatment process or the effect of wastewater

treatment on NPs, is still largely unknown. Hence, it is necessary to investigate how nanomaterials

can be removed from industrial or domestic wastewater and sewage treatment plants, in order to

avoid unintentional releases to natural systems (Theron et al. 2008).

As has been discussed before, sewage sludge contains a major percentage of NPs and metals, which

can be released to the different environmental compartments depending on their final disposal.

There are several disposal alternatives for sludge; for example landfilling, incineration and use as

fertilizer on agricultural land or as a soil improver (Fytianos et al. 1998; Alloway & Jackson 1991).

While in Europe the practice of using sewage sludge as a fertilizer is no longer an alternative, in some

developing countries this is the usual final disposition for this type of waste due to its high content of

organic matter. When using sludge as a fertilizer or disposing it in a landfill, it is expected that in the

long term hazardous substances may be released, percolate to lower layers of the soil and reach

groundwater.

Within the environment NPs may experience aggregation, dissolution, redox reactions, photo-

transformations, among others (Fabrega et al. 2011; Hammes et al. 2013; Lowry, Gregory, et al.

2012). These processes are determined by the characteristics of the NPs (material, size, shape,

surface coating, etc.), the abundance of contaminants and other particles, water chemistry and

hydrodynamic conditions of the water body, and have a great impact on their behavior in the

environment. Aggregation processes reduce the surface area to volume ratio, which in turn affects

7

the transport in porous medium, sedimentation, reactivity, bioavailability, uptake by organisms and

plants, and toxicity (Christian et al. 2008). While the stabilization by surface coatings will maintain

them within the water column, aggregation will likely lead to settling on sediments, reducing the

transport distances and rates, and making benthic organisms a key receptor for nanoparticles.

The environmental risks associated with NPs include bioaccumulation and toxicity. Bioaccumulation

is the process by which chemicals are taken up by organisms from direct exposure or consumption of

food, and conclude in a progressive increase in concentration of the pollutant in the food chain due

to the inability to remove the substance from the body (U.S. Geological Survey). Bioaccumulation will

be determined by the balance between the rate of uptake, i.e. how much they absorb, and rate of

loss and dilution by growth, i.e. how much is excreted from the organism and dilution given by

growth of the organism (Fabrega et al. 2011). In this way the total bioaccumulation will correspond

to the concentration remaining in the organism. Uptake rate varies as a function of bioavailable

concentration and the resultant bioaccumulation will be influenced by the rate and time of exposure.

Both surface area and aggregation of NPs will influence the bioavailability and toxicity of the material

(Levard & Hotze 2012). The size of NPs is directly related with the toxicological effects that these may

cause, hence smaller NPs exhibit larger toxic responses. Likewise, aggregation will affect toxicity by

reducing the dissolution of silver ions and subsequent uptake by organisms, hence altering the

bioavailability of silver in solution.

Nanoparticles may impact the environment on different ways, they may pose a direct risk for biota,

for example toxicity; they can change the bioavailable concentrations of toxins and nutrients;

indirectly affect the ecosystem; and produce changes in the environmental micro-structures

(Christian et al. 2008). Metallic silver is relatively innocuous, however in aqueous environments it

reacts with water and release silver ions (Ag+), which exhibit the antibacterial properties of Ag-NPs

(Ju-Nam & Lead 2008; Kumar et al. 2005). For this reason, their release into wastewater systems may

affect negatively the microbial communities present in the biological processes of the treatment

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plant (Sheng & Liu 2011). Thus, the presence of Ag-NPs could decrease the effectiveness of

contaminant removal in biological treatment stages of the process, and could even pose a problem

for the compliance of standard limits of effluent concentrations. However, the impact of Ag-NPs on

wastewater microorganisms is not yet well understood, and just a scarce amount of reports suggest

that Ag-NPs could significantly inhibit free-floating, planktonic heterotrophic and autotrophic

microorganisms in wastewater. Previous studies (Davies 2003) have shown that biofilms are more

tolerant to antibacterial agents than planktonic bacteria; therefore the effects of Ag-NPs can be

significantly different from those already known. Silver ions bind to negatively charged components

in proteins and nucleic acids, causing structural damage to cell walls, membranes and other cell

structures. Some studies have shown that silver ions inhibit the enzymes of the phosphorous, sulfur

and nitrogen cycles of nitrifying bacteria. Moreover, Ag+ can block DNA transcription; interrupt

bacterial respiration and ATP production and reacts with proteins, inactivating them (Ratte 1999;

Jeon et al. 2003).

2.3 Removal of Ag-NPs during wastewater treatment

Benn et al. (2010) have calculated the amount of silver released per capita from the use of daily

household products (e.g. washing of textiles, use of personal care products and detergents).

Assuming that all released silver reaches the sewage system, one consumer could be responsible for

releasing approximately 470 µg Ag/day. Considering this case scenario, the concentrations of silver in

wastewater biosolids might increase compared to the levels present currently (Benn et al. 2010).

2.3.1 Removal Processes

2.3.1.1 Adsorption mechanisms

Adsorption is the preferential partitioning of substances from one phase, gaseous or aqueous, onto

the surface of a solid substrate (International Adsorption Society). Physical adsorption is caused by

9

electrostatic and Van der Waals forces between the compound and the atoms that compose the

substrate surface. In the case of wastewater, adsorption to organic and inorganic solids is a major

removal mechanism for pollutants in general. For NPs in activated sludge wastewater treatment

plants, it results in an accumulation of pollutants in biosolids to concentrations several orders of

magnitude higher than in influents (Kiser et al. 2010).

Inert and active heterotrophic bacteria and extracellular polymeric substances (EPS) are the main

components of activated sludge, playing a significant role in the adsorption of nanoparticles. Kiser et

al. (2010) have studied the magnitude of adsorption of different types of NPs exposed to wastewater

biomass, and how natural organic matter (NOM), EPS and other factors may affect biosorption. It

was found that even at small concentrations of activated sludge (50 mg/L TSS) high removal

efficiencies (96±1%) can be achieved for Ag-NPs. The adsorption of NPs to biofilm is likely to occur in

two subsequent steps. The first step is the adsorption of nanoparticles to the cell surface or EPS,

most likely driven by electrostatic forces. Afterwards, uptake by the cell could follow, via passive

diffusion or facilitated transport across the cellular membrane.

Also, NPs may be removed from wastewater through adsorption to large debris and suspended solids

present in the influent, and be separated in the preliminary and primary treatments of wastewater,

adding to the efficiency in the removal of nanomaterials during the entire process (Brar et al. 2010).

2.3.1.2 Sulfidation and precipitation

Due to the transformation process and lower solubility of sulfide compounds, sulfidation of Ag-NPs

may be an important detoxification mechanism (Reinsch et al. 2012). The extent of sulfidation

depends on the HS-/Ag ratio and the aggregation state of the nanoparticle, i.e. monodisperse

particles are fully transformed to sulfide compounds, whereas aggregates transform only partly at

the same reaction times. Hence, NPs aggregates will more likely conserve the toxic characteristics of

10

metallic silver. These results suggest that the properties of NPs affect sulfidation products, which in

turn affect microbial growth inhibition.

Sulfidation can occur through two logical mechanisms transforming Ag-NPs to Ag2S. This process may

take place via oxidative dissolution of Ag-NPs followed by precipitation, or by direct oxysulfidation.

The first route described has the disadvantage of intermediate species, silver ions, with potentially

high bioavailability and toxicity. If the concentration of sulfide is high enough compared to Ag-NPs,

this process may result in the generation of a passivating layer of Ag2S around the metallic core of

nanoparticles, thus slowing or preventing the release of silver ions (Reinsch et al. 2012).

The reaction mechanism and effect of environmental conditions over oxysulfidation have been

recently studied by Liu et al. (2011). The results reported confirm that the primary reaction that takes

place is an oxysulfidation, which can be written as a combination of:

4Ag + O2 + 2H2S -> 2Ag2S + 2H2O

4Ag + O2 + 2HS- -> 2Ag2S + 2OH-

The solubility product of silver sulfide is very low (order 10-51), making this compound extremely

stable and thermodynamically favorable in the environment at normal temperatures (Levard & Hotze

2012). It is important to notice that although capping agents may influence the kinetics of the

reaction, it is unlikely that thermodynamically the reaction will be affected by typical capping agents

such as citrate and polyvinylpyrrolidone (PVP).

Kaegi et al. (2011) investigated the fate of NPs in wastewater treatment plants using pilot systems.

The results indicate near-complete sulfidation of Ag-NPs, both in mixed liquor and effluent samples.

These results suggest that the elevated levels of sulfide in non-aerated mixed liquor were responsible

for effective transformation of Ag-NPs to Ag2S. Sulfur is present in wastewater predominantly as

sulfate and originates from urine, which can be transformed to sulfide by sulfate-reducing bacteria

(SRB) (Kaegi et al. 2011). However, in this research no characterization of the properties and toxic

11

potential of sulfidized Ag-NPs were made. The results of Reinsch et al. (2012) demonstrate that the

ability of Ag-NPs to cause Escherichia coli growth inhibition depends on the extent of sulfidation,

which is influenced by the aggregation state and initial size of the NPs. Because the concentration of

sulfide ions in wastewater treatment plants is much higher compared to Ag-NPs it may be suggested

that toxicity to microorganisms due to silver dissolution from the NPs is going to be low. Larger

particles and aggregates show slower oxidation and sulfidation rates. Polydisperse Ag-NPs have

lower rates of sulfidation and higher toxicity potential. As has been stated before, Ag-NPs are mainly

accumulated in biosolids and sludge, which can be applied in soils as a fertilizer. The application of

fully or partially sulfidized Ag-NPs may affect the microbial population of soils, even shift the

organisms present, which in turn can be detrimental to the environment (Reinsch et al. 2012).

Other potential anions present in wastewater that can bind with ionic silver released from NPs

include chloride, carbonates and sulfates (Levard & Hotze 2012). Based on the stability constants

shown in Table 2.1, the reactions with chloride, carbonate and sulfides are more thermodynamically

favored and environmentally relevant than reactions with sulfate.

Table 2.1. Solubility products (Ksp) of silver containing solids (adapted from Levard & Hotze 2012).

Compound Formula Ksp

Silver oxide Ag2O 4.00 x 10-11

Silver carbonate Ag2CO3 8.46 x 10-12

Silver chloride AgCl 1.77 x 10-10

Silver sulfide Ag2S 5.92 x 10-51

Silver sulfate Ag2SO4 1.20 x 10-5

2.3.1.3 Aggregation of Ag-NPs

Aggregation is one of the most important removal processes of NPs. This process can be affected by

environmental factors such as pH, ionic strength, dissolved oxygen (DO) and NOM content (Unrine et

al. 2012). The aggregation of NPs can influence particle properties (i.e. size and morphology) and

transport (i.e. diffusivity) which in turn affect the bioavailability and toxicity.

12

El Badawy et al. (2010) have studied the stability of Ag-NPs by evaluating the changes in zeta-

potential as a function of pH, ionic strength and electrolyte species. From their results is possible to

conclude that in environments with acidic conditions or high ionic strength, especially those that

contains divalent ions, is more likely that Ag-NPs will aggregate and settle. Furthermore, it was

concluded that Ag-NPs are potentially unstable in environments such as landfills, wastewater, soils,

surface and groundwater. Another factor that may influence the aggregation of Ag-NPs is the

presence of DO in the aqueous environment (Zhang et al. 2011). When DO is present, Ag-NPs not

only aggregate but also release silver ions due to oxidation, which will induce other physicochemical

processes, e.g. complexation with NOM and electrolytes, precipitation and dissolution. On the other

hand, conditions with low redox potential such as anoxic and anaerobic environments inhibit the

release of silver ions. The results of Zhang et al. (2011) suggest that in presence of DO the

aggregation extent is higher than in absence of it. Furthermore, the presence of chloride ions in the

medium may induce the aggregation of Ag-NPs by the formation of crystalline compounds (Botasini

& Méndez 2013).

The capping agent of NPs is used to protect them from aggregation and further settling, by providing

colloidal stability through electrostatic or steric repulsion (El Badawy et al. 2010). In the synthesis of

Ag-NPs the most used capping agents are citrate, sodium borohydrate and PVP. The results of

previous studies (El Badawy et al. 2010) suggest that the use of capping agents can dramatically

influence the aggregation of Ag-NPs.

2.3.1.4 Bioaccumulation by microorganisms and plant uptake

Some living organisms cells are surrounded by semipermeable walls, which allow the entrance of

particles smaller than their pore size, hence Ag-NPs, and NPs in general, will pass through the cell

wall and reach the plasma membrane (Navarro et al. 2008; Fabrega et al. 2011). This plasma

membrane is a complex system with known mechanisms and structures, such as protein carriers and

13

pores that permit the translocation of material across the membrane. Two possibilities of NPs

transport through the membrane are probable, endocytosis and association with surface. During the

first option, endocytosis, the plasma membrane forms a cavity-like structure that surrounds the NP

and transports it into the cell. The association of Ag-NPs with the surface implies a further release of

silver ions within the surface layers, creating environments with high silver concentration, which

leads to a rapid uptake of the ion.

The rates of uptake and bioaccumulation are influenced by characteristics of the nanoparticles (size,

shape, surface charge, chemical composition and area, solubility and aggregation state of the

particle) and environmental conditions (pH, ionic strength, composition, NOM and temperature),

parameters that will affect the aggregation and stability of NPs (Fabrega et al. 2011). While the

uptake rate varies as a function of the bioavailable concentration, bioaccumulation will be affected

by the time and concentration of exposure.

On the other hand, plants can easily take up species of metals dissolved in solution, either as ionic or

chelated forms. The amount of metals that can be taken up depends on the plant species and the

stage of development of it. Absorption by the root system is the main pathway by which metals can

be taken up by plants (Kabata-Pendias & Pendias 2001). This absorption can be either passive

(nonmetabollic) or active (metabolic), depending on whether the transport is in favor or against the

chemical gradient.

In the case of Ag-NPs, it has been already shown that plants can sorb and take up silver, both in ionic

form or NPs forms, even though the mechanism is not known (Bone et al. 2012). The degree of

aggregation of NPs is an important factor that may influence the available amount of silver that

plants can take up. A decrease in aggregation is generally related with an increase in bioavailability,

due to a higher concentration of NPs in the water column in comparison with particles settled on

sediments. Harris & Bali (2007) have studied the plant metal uptake rate, and concluded that at

14

higher exposure time and concentration, the amount of metal that plants are able to accumulate will

increase.

Wetland systems are generally a sink for metals, due to the anoxic conditions prevailing in parts of

the system (Weis & Weis 2004). The bioavailability of metals is low compared to other systems

where soils are mostly oxidized. Changes in pH and redox potential influence the speciation of metals

and their solubility, which can result in an increase or decrease of concentration in the water column,

and subsequent increase or decrease in uptake by plants. The different species that can be found

have different availabilities: water soluble and exchangeable metals are the most available, while

metals precipitated as inorganic compounds, complexes and adsorbed to hydrous oxides are

potentially available, and on the other hand metals precipitated as insoluble sulfides or bound to

crystalline lattice of minerals are essentially unavailable. Even though wetland sediments act as sinks

for pollutants because of their anoxic conditions, the plants present can become a source of metals.

Plants can oxidize the sediments in vicinity of the root system, making metals present in different

forms available for other reactions and for plant uptake.

Wetland plants can also translocate metals from root to aerial tissue, accumulating them in leaves

and stems (Weis & Weis 2004). The degree of translocation will depend on the plant species, the

particular metal and environmental conditions. Metals can be taken up by the roots, and transported

to above-ground tissues, from which they can be excreted. The transport of metals within the plant

and their possible release from leaf tissue are important steps in the study of metal fluxes in wetland

ecosystems.

2.3.2 Conventional wastewater treatment plants

Municipal wastewater treatment plants (WWTP) are designed to remove undesirable components

from the influent, such as organic materials, nutrients, solids and pathogens (Vesilind et al. 2010).

The typical WWTP is divided into five main processes, as shown in Figure 2.1:

15

1) Preliminary treatment for the removal of large solids, through the use of screens and grit or

sand chambers.

2) Primary treatment to remove suspended solids through sedimentation, in settling tanks or

clarifiers. As a consequence of the removal of raw sludge, 60% of the solids, 30% of the

demand of oxygen and 20% of phosphorus are removed in the primary treatment.

3) Secondary treatment for the degradation of organic compounds. Almost all secondary

treatment methods use microbial action to reduce the energy level of the waste (oxygen

demand), with differences in how the contact of wastewater and microorganisms is

established. The microorganisms use the energy and carbon by decomposing the organic

compounds to CO2, H2O and other stable compounds, and the production of new

microorganisms.

4) Tertiary or advanced treatment comprising any form of polishing or clean-up processes, like

the removal of nutrients, rapid sand filters, oxidation ponds and activated carbon adsorption.

Also, due to the expense and complexity of these treatments, natural systems have been

developed such as different types of land treatment and constructed wetlands.

5) Solids treatment and disposal, which include collection, stabilization and disposal of the

solids removed from previous stages. The objective of sludge stabilization is the reduction

sludge odor and putrescence and the presence of pathogens. Finally, dewatered sludge

needs to be disposed, and depending on legal and technical constrains this can be achieved

through incineration, landfill or use as fertilizer.

16

Figure 2.1. A typical wastewater treatment plant including preliminary treatment, primary treatment,

secondary treatment, tertiary treatment and solids treatment and disposal (Brar et al. 2010).

As previously stated, the majority of nanoparticles will reach WWT systems. During the wastewater

treatment NPs can be removed from the water through several mechanisms in the subsequent

stages of the process (Brar et al. 2010). For instance, during the pretreatment phase the main

removal mechanism is the adsorption of NPs to debris or other large particles. However, it is possible

that NPs find their way through wastewater to the primary treatment phase. At this stage, as in most

cases coagulants are used to enhance the sedimentation of particles. This practice may lead to the

adsorption of NPs and their further settling in wastewater sludge. During secondary treatment there

is a possibility that NPs will adhere to microbial cell surfaces and associated EPS, or be taken up by

microorganisms. The processes and characteristics of media and NPs that determine the transport of

NPs toward microorganisms comprise diffusion, gravitational settling and agglomeration.

Biodegradation of NPs may result in their breakdown as typically seen in biodegradation of organic

molecules, or could produce changes in the physical structure or surface characteristics of the

material, as is the case with metallic NPs, which undergo biodegradation of the organic coating

surrounding the metal core.

17

Kaegi et al. (2013) performed a batch experiment using silver and gold NPs, with different size and

coating. In all cases, efficiencies of ~99% for the removal of NPs were found, suggesting their quick

incorporation or attachment to flocs of the activated sludge. From these results, it can be assumed

that the majority of NPs in the treatment plant will be incorporated into sludge and be removed from

the effluent stream. Nevertheless, a small fraction can be discharged to surface water, corresponding

to NPs associated with sludge loss through the effluent.

Wang et al. (2003) studied the interactions of silver with wastewater constituents. Their results have

shown that most silver can be removed through precipitation with chloride and adsorption by sludge

particulates. Other removal mechanisms include the complexation and precipitation of silver with

inorganic and organic materials such as chloride, sulfide, thiosulfate and dissolved organic carbon

(DOC). Some studies (Kim et al. 2010) have reported that the majority of Ag-NPs that enter the WWT

are sulfidized and accumulated in the form of AgS2 in sludge and biosolids. Li & Hartmann (2013)

analyzed the concentration of silver in the different stages of the WWT process in several treatment

plants in Germany, and reported that mechanical separation phases play an important role in the

removal of silver in general and also Ag-NPs. This fact has been attributed to the attachment to larger

particles like suspended organic matter. Also, it has been proven that the remaining silver can be

reduced in the biological treatment by 72-99%. From the wastewater that enters the treatment

plant, at least 95% of Ag-NPs are efficiently removed from the effluent (Li et al. 2013; Kaegi et al.

2011).

2.3.3 Constructed wetlands for wastewater treatment

The term wetland encompasses a broad range of environments, characterized by the wet conditions

present, including marshes, swamps, wet meadows, floodplains and riparian wetlands along stream

channels (Davis n.d.). Wetlands are complex systems that comprise water, plants, animals,

microorganisms and the natural surroundings. Recent concerns over wetland losses have initiated

the creation of constructed wetlands (CW) that emulate the functions and values of the natural ones,

18

in terms of physical, chemical and biological processes and characteristics (Wetlands International -

Malaysia Office 2003). A CW is a shallow basin filled with a substrate, usually gravel or soil, and

planted with helophytes tolerant to saturated conditions (Davis n.d.). In most wetlands, the

hydrological conditions are such that oxygen poor conditions are achieved in the substrate, limiting

the vegetation that can grow to certain species adapted to low-oxygen environments. Water is

introduced at one end of the wetland and it flows over the surface or through the substrate until the

outlet. Some CW are specifically created for WWT to provide secondary or tertiary treatment of

domestic, industrial and agricultural wastewater (Wetlands International - Malaysia Office 2003).

These systems are collectively termed “Constructed Wetland Treatment System” (CWTS). During the

1970s and 1980s wetland systems were used nearly exclusively for the treatment of domestic or

municipal wastewater (Vymazal 2005). However since the 1990s, they have been used in various

types of wastewater including landfill leachate, runoff, food processing, industrial, agriculture farms,

mine drainage or sludge dewatering.

Natural and constructed wetlands represent a low-cost alternative technology for WWT, offering an

economically attractive and energy-efficient way of providing WWT of high quality (Davis n.d.;

Wetlands International - Malaysia Office 2003). Wetlands are characterized by a high organic matter

accumulation due to a high rate of primary production and low rate of decomposition (Kivaisi 2001).

Wetlands have been found to be effective in the treatment of biochemical oxygen demand (BOD),

total suspended solids (TSS), nutrients such as nitrogen and phosphorous, and the reduction of

metals, organic pollutants and pathogens. Also, wetlands are constructed in order to compensate

habitat loss due to agriculture and urban development, water quality improvement, flood control,

production of food and fiber, recreation and landscape enhancement. Data compiled by Verhoeven

& Meuleman (1999) show that the performance of wetlands for WWT depends on the loading rate

and specific hydrological and ecological characteristics of the wetland. In the case of organic matter

and bacterial pollution removals of 80-99% can be achieved, for nutrients values are more variable

and lower. Nitrogen removal was higher than 50% and phosphorous was lower.

19

The main mechanisms by which pollutants are removed include biological processes and

physicochemical processes such as filtration and sedimentation of suspended solids; adsorption in

the surface of plants, substrate, sediment and litter; precipitation and chemical transformation

(Kivaisi 2001; Davis n.d.). Microbial degradation plays an important role in the removal of soluble and

colloidal organic matter present in the wastewater. This degradation process occurs when dissolved

organic matter (DOM) is transported into the biofilms attached to plant stems, roots and soil or

media by diffusion. Also, processes of plant and microorganisms uptake and transformation can

occur in wetland systems.

The basic classification of wetlands types is based in the macrophytic growth: emergent, submerged,

free floating and rooted with floating leaves (Vymazal 2005). Also, they can be classified based on the

water flow regime: surface flow (SF), sub-surface flow (SSF), either vertical or horizontal. Figure 2.2

shows a general scheme with the classification and operation of different types of CWTS. Surface

flow systems are used mainly for municipal wastewater treatment with large water flows for nutrient

polishing, and the water flows on the surface of the wetland. On the other hand, in SSF systems

water flows from one end to the other through a permeable substrate, usually a mix of gravel, rock

and soil. The substrate supports the growth of plants and microorganisms. Sub-surface flow systems

are best suited for wastewaters with low solids concentrations and uniform flow conditions (Davis

n.d.). These systems can be smaller compared with SF systems, however they are more expensive to

construct and maintain. In wetlands with vertical flow, also called infiltration wetlands, the

wastewater flows vertically through highly permeable sediment and is collected in drains at the

bottom (Verhoeven & Meuleman 1999).

20

Figure 2.2. Classification of constructed wetlands for wastewater treatment and scheme of the operation of

free floating plants, surface flow and sub-surface - horizontal and vertical - flow constructed wetlands (top to

bottom) (Vymazal 2007).

Some advantages of this kind of wastewater treatment systems have been listed by Wetlands

International - Malaysia Office (2003) including:

1) a better aesthetical looking site compared with conventional WWTP;

2) lower cost overall, due to the use of renewable energy, like solar and kinetic energy, plants

and microorganisms, that are an important component in the treatment process, and lower

complexity compared with other treatment methods;

3) the systems can tolerate a varying level of contaminants and volumes of water;

4) they could be used to clean up polluted rivers and other water bodies;

5) it can also provide habitat for wetland species, and could become an attractive tourism

destination.

21

Due to these advantages, especially the low cost and maintenance, CWTS have a strong potential for

application in developing countries and small rural areas (Kivaisi 2001). However, it is important to

take into account some important limitations for these systems, such as the larger area needed

compared to conventional processes. Moreover, efficiencies can vary seasonally in response to

environmental conditions as drought and rainfall, making these treatment method less consistent in

terms of performance, and they can be sensitive to toxic compounds due to the natural character of

their components (Davis n.d.).

Wetlands are composed from a variety of components that influence removal processes of pollutants

and nutrients. For the design of CW, the selection of substrates is of utmost importance. Substrates

include soil, sand, gravel, rock and organic materials. Substrates may remove wastewater pollutants

by physicochemical means, such as adsorption, ion exchange, precipitation and complexation; and

provide a suitable medium for plant growth, infiltration and movement of wastewater (Wetlands

International - Malaysia Office 2003; Davis n.d.). The chemical composition, hydraulic permeability

and capacity to adsorb nutrients and contaminants determine the efficiency and processes that will

take place in the wetland. When selecting substrates for wetland construction some properties of

them have to be considered, such as cation exchange capacity (CEC), pH, electrical conductivity (EC),

redox potential, texture and soil organic matter. These characteristics affect the processes that are

occurring in the wetland system. Also important in CW are plants and microorganisms. Wetland

plants have an important role in the removal and retention of nutrients and provide a great surface

area for the attachment and growth of microbes. They also slow down the water flow enhancing the

gravitational settling, thus increasing water transparency (Wetlands International - Malaysia Office

2003). Microbial activities transforms organic and inorganic substances into innocuous or insoluble

substances, alters redox conditions affecting processes in wetlands, and are involved in the recycling

of nutrients (Davis n.d.).

22

Kröpfelová et al. (2009) have studied the removal of trace elements in wetland systems that may be

a starting point to identify the different processes that can affect the fate of different metallic NPs.

The most important factors that influence mobility of trace elements in wetland systems are sulfide

and iron/manganese hydrous oxide formation and dissolution, which are directly related to the redox

conditions in the system. Under reducing conditions many trace elements react with hydrogen

sulfide to form insoluble metal sulfides. However, under oxidized conditions the oxidation of sulfides

to sulfates will release these metals into the water. The results obtained by Kröpfelová et al. (2009)

indicated that trace elements are retained with various efficiencies ranging from 90% for aluminum

to 57% for arsenic.

Lowry, Espinasse, et al. (2012) have described the transformation of PVP-coated Ag-NPs in a

freshwater wetland mesocosm by determining their partitioning behavior, the speciation and

bioavailability of silver to plants and aquatic species. One of the main factors that may affect the fate,

distribution and speciation of NPs in natural systems is the media through which they are introduced

in the wetland system. For example, Ag-NPs can enter aquatic environments through WWTP effluent.

In this case Ag-NPs will likely be associated with other natural colloids and settle in aquatic sediment.

On the other hand, they can enter terrestrial ecosystems by the application of biosolids containing

NPs. Here, they can be mobilized and transported to receiving waters, and be accumulated in

subaquatic sediments. Two routes of introduction were used in order to study the long term fate of

citrate-coated Ag-NPs: water column and terrestrial application. The results show that Ag-NPs rapidly

settle from the water column, as evidenced by the drop of total silver concentration in the aqueous

phase, as illustrated in Table 2.2 This table shows the silver mass recovered in the different

compartments for both routes of introduction of Ag-NPs. A redox profile was found to occur in the

sediments, i.e. an oxic layer in the upper layer and anoxic conditions in deeper layers, which was

found to influence the oxygen penetration and oxidation state of metals. The majority of the silver

remained in the location where it was dosed, but a continuous transfer between the terrestrial and

aquatic phases occurred during the 18 months that the experiment was performed. In general, silver

23

is primarily present in the upper layer of sediment and terrestrial soils, and a small mass of silver was

found in the plants biomass at the end of the study.

Table 2.2. Estimated distribution of mass of silver recovered in mesocosms dosed in either the terrestrial

compartment or the water column (Lowry, Espinasse, et al. 2012).

Compartment Location of dose

Terrestrial soil Water column

Water column <0.03% <0.05%

Terrestrial soil 58% 7%

Subaquatic sediment 11% 60%

Plants 3% 0.2%

Total recovered 76% 68%

After the experiment, i.e. after 18 months, the Ag-NPs were partially oxidized and sulfidized. Particles

recovered from the terrestrial soils contained Ag2S and Ag0, 52% and 47% respectively, whereas the

Ag-NPs recovered from the subaquatic sediment were more oxidized than those in the terrestrial

compartment, where only 18% of Ag0 remained after aging. Sulfidation was found to be partial even

after 18 months, indicating that the process occurs much slower than observed in laboratory

experiments. This may be caused by a lower concentration of sulfide or Ag-NPs, and the competition

with other metals present in natural systems. The sulfidation rate can also be hindered by the

concentration of DO required in the oxidation process.

The partitioning behavior of Ag-NPs between solid and aqueous phase indicates that they can move

between compartments. This suggests that runoff and flooding events can mobilize soil particles and

Ag-NPs attached to them, providing a potential release for NPs enter surface waters. Accumulation

of silver in plants and animals suggests that a portion of the silver remains bioavailable after the

gravitational settling into aquatic sediment and partial sulfidation.

Nevertheless, the research on fate of NPs in wetland systems is still new, and more work needs to be

done on this topic in order to better understand the different processes involved in their removal

and avoid unintentional releases to natural ecosystems.

24

3. Materials and Methods

3.1 Synthetic wastewater

Synthetic wastewater was prepared according to Weber et al. (2011). Table 3.1 shows the final

concentrations of the different constituents. A nutrient solution A was prepared with concentration

1000 time higher than necessary. Nutrient solution A was diluted 1000 times for each wastewater

preparation, while solution B was freshly prepared for each experiment. The final concentration of

COD in the synthetic wastewater was approximately 500 mg/l, and the proportion of nutrients was

COD:N:P as 100:5:1. All the compounds used are commercially available chemical reagents and the

molasses were purchased in a local store.

Table 3.1 Concentration of the constituents of synthetic wastewater used

in the experiments.

Nutrients solution A

Compound Concentration mg/L

FeNaEDTA 9.175

H3BO3 0.715

MnCl2*4H2O 0.4525

ZnSO4*7H2O 0.055

CuSO4 0.0125

(NH4)6Mo7O24*4H2O 0.005

Nutrients solution B

Compound Concentration mg/L

NH4H2PO4 28.75

KNO3 151.5

Ca(NO3)2*4H2O 236

MgSO4*7H2O 123.25

Urea 49

NH4H2PO4 18.5

Molasses 1000

25

3.2 Characterization of nanoparticles and substrates

3.2.1 Characterization of AgNP solution

An aqueous suspension of Ag-NPs stabilized with citrate, distributed by PlasmaChem GmbH, was

used. The average size reported by the producer was 10 nm and the concentration of the solution

was 0.1 mg/ml Ag-NPs.

The size distribution of the NPs used was determined using a Photon Correlation Spectrometer (PCS

100M Malvern). The sample was analyzed in triplicate and the mean particle size was determined

based on intensity and number.

3.2.2 Characterization of substrates

Sand, gravel and zeolite substrates were used as solid components of wetland systems. Rhine gravel

was selected with a size of 8-16 mm and zeolite size was 5-10 mm. The substrates were washed with

tap water prior to use in order to remove the fine particles present.

3.2.2.1 pH

For the measurement of pH-H2O 10 g of each substrate were mixed with 50 ml of water. The

suspension was allowed to equilibrate for 24 hours. The supernatant was measured using a pH

electrode (Orion 520A). Analysis was done in triplicate for each substrate.

3.2.2.2 Electrical conductivity

The electrical conductivity was determined in each substrate by transferring 10 g of substrate into a

250 ml Erlenmeyer flask, adding 50 ml of distilled water and shaking the suspensions during 30 min

on a shaking plate. Afterwards, the samples were filtered through a filter paper and the EC of the

filtrate was measured with a conductivity probe (WTW LF537). Analysis was conducted in triplicate.

26

3.2.2.3 Organic matter content

The organic matter content (%) was determined by transferring 2 g of substrate into small crucibles

and igniting them in an oven at 550°C for 24 hours. The organic matter content was calculated as the

difference in weight before and after ignition.

3.2.2.4 Cation exchange capacity (CEC)

In order to assess the capacity of each substrate to exchange cations, a saturation of the adsorption

sites with ammonium ions (NH4+) and subsequent release of these ions with 1M KCl solution was

performed. Depending on the substrate being analyzed different sample amounts were used. For

gravel 60 g were used, for zeolite 20 g, and for sand 10 g. These amounts of substrate were

transferred to percolation tubes. After percolation of 150 ml 1M ammonium acetate (NH4OAc),

300 ml denatured ethanol (C2H5OH) was added to wash away the excess NH4+ ions. Afterwards,

250 ml 1M KCl solution was added to remove the exchangeable NH4+ ions. The final percolate was

then captured in a 250 ml volumetric flask. Afterwards, 50 ml of the KCl extract was transferred to a

distillation flask and 0.1 mg MgO were added. The NH3 formed was captured by steam distillation as

NH4+ in an Erlenmeyer flask that contained 20 ml of 2% boric acid. This final solution was titrated

with 0.01 M HCl using a Methrohm 718 STAT Titrino system until the color shifted to pink.

3.2.2.5 Determination of trace elements

Concentrations of Na, K, Mg, Ca, Zn, Al, Cd, Cr, Cu, Fe, Mn, Ni and Pb in each substrate were

determined via ICP-OES after digestion with aqua regia. Two grams of substrate were transferred to

an Erlenmeyer flask, to which 5 ml H2O, 15 ml HCl and 5 ml HNO3 were added. The suspension was

left overnight. Afterwards, the samples were boiled on a heating plate at 150°C during 2 hours. After

cooling, the samples were filtered, transferred to a 100 ml flask and subsequently diluted to 100 ml

27

with deionized water. This solution was analyzed via ICP-OES (Vista-MPX CCD Simultaneous ICP-OES).

In the same digest, Ag was analyzed using ICP-MS (Perkin Elmer SCIEX ELAN DRC-e).

3.3 Adsorption experiments

Adsorption experiments were performed to assess the adsorption capabilities of the different

components of wetland systems (solid matrix, biofilm and plant roots). Each experiment was

performed in triplicate according to the scheme in Figure 3.1. In each experimental batch a blank and

a sample of known concentration prepared with Milli-Q water were analyzed.

Figure 3.1 Scheme of adsorption experiments performed with solid

substrates, and subsequent sample preparation and analysis.

28

3.3.1 Preliminary test to evaluate the effect of centrifugation speed

In order to evaluate the aggregation and precipitation of silver compounds in wastewater during 24

hours, centrifugation experiments were performed. Samples prepared with Milli-Q water and 100

µg/l Ag-NPs were analyzed at different centrifugation speed. 100 ml of sample were transferred to a

300 ml Erlenmeyer flask, from which a 10 ml sample (Sample I) was taken to determine the initial

concentration of Ag-NPs in the solution. These samples were centrifuged at 0, 500 or 1000 rpm for

10 minutes, acidified and stored for further analysis. The remaining 90 ml were shaken during 24

hours on a shaking plate. Later, a 10 ml sample (Sample II) was centrifuged and acidified. All samples,

before and after the 24 hours mixing period, were digested in a microwave oven (Mars 5) after

previous addition of 2 ml HNO3 (65%) to 5 ml of sample. Digested samples were diluted to 10 ml and

analyzed via ICP-MS (Perkin Elmer SCIEX ELAN DRC-e) to determine the Ag concentration.

3.3.2 Preliminary test to evaluate the effect of wastewater strength

To assess the effect that wastewater strength could have on the fate of the Ag-NPs in solution in

terms of the different reaction that can form silver compounds that precipitate in solution, Ag-NPs

were spiked in solutions containing different doses of synthetic wastewater (10, 50 and 100% of

wastewater). These solutions were transferred to 300 ml Erlenmeyer flasks, from which a 10 ml

sample (Sample I) was taken to determine the initial concentration of Ag-NPs in the solution. These

samples were centrifuged at 500 rpm during 10 minutes and acidified afterwards. The remaining

sample (90 ml) was shaken during 24 hours on a shaking plate and a 10 ml sample (Sample II) was

taken at the end of the mixing period. An acid digestion in a microwave oven (Mars 5) using 2 ml of

HNO3 65% was carried out on 5 ml of sample. The samples were diluted to 10 ml and analyzed via

ICP-MS (Perkin Elmer SCIEX ELAN DRC-e) to determine Ag concentration.

29

3.3.3 Adsorption on solid substrates

In order to assess the adsorption potential of the different substrates used in wetland systems,

adsorption experiments were performed. Solutions (100 ml) of 0, 100 and 1000 µg/l Ag-NPs prepared

in a background of 100% wastewater were transferred to 300 ml Erlenmeyer flasks. A 10 ml

subsample (Sample I) was taken and acidified for further analysis. The remaining solution (90 ml) was

mixed with 30 g of substrate (gravel, sand or zeolite) and shaken during 24 hours on a shaking plate.

Afterwards, a 10 ml sample (Sample II) was taken. Five ml of samples I and II were digested with 2 ml

of HNO3 65% in a microwave oven (Mars 5), subsequently diluted to 10 ml and analyzed via ICP-MS

(Perkin Elmer SCIEX ELAN DRC-e) for Ag determination.

3.3.4 Effect of biofilm formation

To study the adsorption of Ag-NPs on biofilm, adsorption tests with active and passive biofilm were

performed using gravel taken from a pilot setup wetland. Furthermore, adsorption experiments with

no biofilm were performed on fresh gravel substrate. A sample of gravel from the pilot setup was

taken and half of it was dried at 40°C for 24 hours, in order to passivate the biofilm on the surface,

decreasing its metabolic activity. The adsorption experiment was carried out with dried (passive

biofilm) gravel samples, on fresh (active biofilm) gravel samples from the wetland setup system and

clean (no biofilm) gravel. Hundred ml solutions of wastewater spiked with 100 µg/l Ag-NPs were

transferred to Erlenmeyer flasks and 10 ml subsamples (Sample I) were taken and acidified in order

to determine the initial Ag concentration added. The remaining solution (90 ml) was mixed with 30 g

of gravel and shaken for 24 hours. Afterwards, a 10 ml sample was taken (Sample II). Five ml of

samples I and II were digested with 2 ml of HNO3 in a microwave oven (Mars 5), and finally diluted to

10 ml and analyzed via ICP-MS (Perkin Elmer SCIEX ELAN DRC-e) for Ag concentration.

Additionally, organic matter content (3.2.2.3) analyses were performed, in order to calculate the

amount of dry matter and biofilm (organic matter) present in the substrate.

30

3.4 Adsorption to plant material and uptake by wetland plants

In order to assess the amount of Ag-NPs that wetland plant (Phragmites australis) roots can adsorb

and absorb, a 4- weeks experiment was set up and a mass balance was calculated. From a pilot

wetland setup, three active and three passive plants were taken and cleaned with tap water. Strong

looking, green plants were selected as active wetland plants and plants with drying aboveground

tissues were used as passive plants. Active wetland plants were composed of roots and aboveground

tissues, while passive wetland plants were just root material. Each plant root was weighted and

transferred to a 500 ml Erlenmeyer flask. All flasks were placed under a growth lamp. Each week a

solution of synthetic wastewater spiked with a known concentration of Ag-NPs was prepared and 10

ml subsamples (Sample I) were taken and acidified for further analysis. The remaining 190 ml were

added to the flask containing the roots. During each subsequent week, the volume of water was kept

constant by adding demineralized water. Afterwards the final volume of water was measured and a

10 ml sample (Sample II) was taken. The rest of the water phase was removed and new solution was

added in the system. Samples I and II (10 ml) were digested in a microwave oven (Mars 5) after

adding 2 ml HNO3 65%. Afterwards, samples were diluted to 20 ml and the Ag concentration was

measured via ICP-MS (Perkin Elmer SCIEX ELAN DRC-e).

At the end of the 4 weeks, after removing the wastewater, different analyses were performed in

order to determine the amount of Ag-NPs adsorbed to the glass walls of the Erlenmeyer flask, roots

and the amount that was incorporated in the plant material, roots and aboveground material. To

determine the adsorption to the glass, 100 ml HNO3 5% and 1 ml H2O2 were added to the empty flask

and shaken during 24 hours. A 20 ml sample was taken and digested in a microwave oven (Mars 5) by

adding 4 ml HNO3 65%. The samples were analyzed using ICP-MS to determine the silver content.

For the determination of the amount of Ag adsorbed to the roots, these were transferred to a

Erlenmeyer flask and 50 ml of Milli-Q water was added, the flask containing the plant roots was

sonicated during 10 minutes to achieve desorption of loosely adsorbed Ag, and a 10 ml sample was

31

taken and acidified. Afterwards, 50 ml 1% HNO3 were added and sonicated during 5 minutes to

achieve a more complete desorption of Ag-NPs from the roots. Again a 10 ml sample was taken and

acidified. Both the water and acid extracts were digested in a microwave oven and the silver

concentration was determined by ICP-MS.

To measure the amount of Ag taken up by the plant material, roots and above ground tissues were

treated separately. Plant material was divided into roots and aboveground tissues (new and old), and

cut in small pieces. Old aboveground tissues were selected as those already present in the active

plants at the start of the experiment, while the new tissues started to grow during the 4 weeks

period, in both active and passive plants. Samples were dried during 24 hours at 105°C in an oven,

and their weight was measured. After drying, roots and leaves were grinded and 0.30 g were taken

for analysis. An acid digestion was performed by adding 5 ml HNO3 65% and 1 ml H2O2 and left

overnight. The samples were further digested using a microwave oven and the silver concentration

was determine via ICP-MS, according to Xiao (2013).

The mass balance of Ag in the water, root and above ground tissue was calculated by weighing the

total amount of Ag-NPs added to the systems against the silver concentrations in the different

analyzed compartments, i.e. the weekly water samples, the adsorption to glass, the adsorption to

roots, and the absorption in roots and translocation to new and old aboveground tissues. Figure 3.2

presents the different compounds taken into account to create the mass balance.

32

Figure 3.2 General scheme of the different compounds taken into account to construct the mass balance of

Ag.

3.5 Statistical analysis

In order to assess whether any significant differences were present between various groups,

statistical test were performed in the software package IBM SPSS Statistics 22. The tests used were

One-way ANOVA, Welch-ANOVA for groups that failed to fulfill all assumptions and Tukey test to

conduct multiple comparisons.

33

4. Results

4.1 Characterization of nanoparticles and substrates

4.1.1 Characterization of Ag-NPs solution

In order to characterize the NP suspension of PlasmaChem GmbH used in the experiments, the

nanoparticle size distribution was determined using a Photon Correlation Spectrometer (PCS)

Malvern, by the research group Paint at Ghent University. Figure 4.1 shows the results in terms

of intensity and number. When based on intensity, two peaks can be observed: one at

90.9 ± 7.9 nm and one at 15.8 ± 0.9 nm. When based on the number, only one peak at particle

size of 11.1 ± 0.2 nm is observed.

Figure 4.1 Size distribution of silver nanoparticles solution analyzed by Photon Correlation

Spectrometry. Left: Size distribution based on intensity; Right: Size distribution based on number.

4.1.2 Characterization of substrates

The different substrates, sand, gravel and zeolite, were analyzed in order to determine the

main physicochemical properties that they possess and how these properties could affect the

fate of nanoparticles in wetland systems containing these substrates. Table 4.1 shows the pH,

EC, CEC and organic matter content of these substrates. In general, the substrates have neutral

Size distribution(s)

5 10 50 100

Diameter (nm)

10

20

% in

cla

ss

Size distribution(s)

5 10 50 100

Diameter (nm)

10

20

30

40

% in

cla

ss

34

to alkaline pH values. Gravel and zeolite have a similar EC, whereas sand shows a higher value,

over 1000 µS/cm. Moreover, the CEC of sand and gravel was lower than 1 meq/100 g

substrate, while the CEC value of zeolite was 11.36 meq/100 g substrate. Zeolites have the

highest organic matter content, as determined by loss on ignition and gravel the lowest.

Table 4.1 Main physicochemical characteristics of different wetland substrates (average ± SD, n=3).

Sand Gravel Zeolite

pH-H2O 7.9 ± 0.1 8.2 ± 0.1 7.9 ± 0.1

Conductivity (µS/cm) 1284 ± 67 30.8 ± 4.6 34.3 ± 3.2

Cation exchange capacity

(meq/100 g substrate) 0.81 ± 0.09 0.43 ± 0.13 11.36 ± 1.00

Organic matter content (%) 2.32 ± 0.45 1.37 ± 0.31 10.03 ± 0.76

In addition to physicochemical properties, the concentrations of major and trace elements and

silver were analyzed for the different substrates (Table 4.2). Zeolite contains clearly a higher

concentration of K, Al, Pb, Zn and Ag than sand and gravel. On the other hand, a higher

concentration of Ca was found in gravel (371.4 ± 3.6 mg/g). The three substrates all have

concentrations of Cd and Cr below the detection limit of the method, which is also the case for

Cu and Ag in sand and Ag in gravel.

Table 4.2 Trace metals concentrations in sand, gravel and zeolite determined via ICP-OES after aqua

regia digestion (average ± SD, n=3). (<D.L.: Lower than Detection Limit)

Sand Gravel Zeolite

Ca mg/g 26.4 ± 5.8 371.4 ± 3.6 19.5 ± 1.0

Mg mg/g 0.51 ± 0.03 2.62 ± 0.48 4.45 ± 0.11

K mg/g 0.18 ±0.01 0.14 ± 0.09 16.41 ± 1.35

Na mg/g 1.23 ± 0.10 0.12 ± 0.01 1.44 ± 0.06

Al mg/g 0.47 ± 0.03 0.38 ± 0.22 27.2 ± 2.2

Cd mg/g <D.L. <D.L. <D.L.

Cr mg/g <D.L. <D.L. <D.L.

Cu µg/g <D.L. 0.909 ± 0.38 1.632 ± 0.17

Fe mg/g 2.01 ± 0.19 1.31 ± 0.40 3.35 ± 0.38

Mn µg/g 39.6 ± 0.6 85.8 ± 23.2 77.1 ± 29.2

Ni µg/g 1.260 ± 0.19 2.341 ± 0.3 2.495 ± 1.161

Pb µg/g 2.18 ± 0.63 0.82 ± 0.82 25.93 ± 2.12

Zn µg/g 5.26 ± 0.27 6.60 ± 2.92 26.31 ± 3.36

Ag ug/g <D.L. <D.L. 0.359 ± 0.041

35

4.2 Effect of centrifugation speed and wastewater strength

In order to assess the loss of Ag-NPs during mixing experiments of 24 hours at different

synthetic wastewater content in solution and subsequent centrifugation speed, a preliminary

mixing experiment was executed. Figure 4.2 shows the difference in concentration of Ag

(Sample I and Sample II, before and after the experiment) in the wastewater phase after 24

hours and 10 minutes centrifugation at different speeds: 0, 500 and 1000 rpm. No significant

differences (p=0.673) are observed when centrifuging the spiked wastewater solutions at

these different speeds. This information was used when planning the adsorption experiments

executed with sand, gravel, zeolites and biofilm. Due to the lack of effect of centrifugation and

its speed on the concentration in the water, the samples in further experiments were not

centrifuged prior to analysis.

Figure 4.2 Effect of the centrifugation speed on the concentration of Ag in the

supernatant after 24 hours of mixing followed by 10 minutes of

centrifugation.

Additionally, Ag-NPs were dissolved in solutions containing different amounts of synthetic

wastewater (0, 10, 50 and 100%) and subjected to a 24-hours mixing experiment, to evaluate

the effect of wastewater content on removal of Ag from the water phase in absence of solid

substrate (Figure 4.3). The percentage of wastewater in solution does not significantly affect

the Ag concentration (p=0.547) in the water phase. This suggests that the components of

0

5

10

15

20

25

0 500 1000

Dif

fere

nce

in A

g co

nce

ntr

ati

on

(u

g/L)

Centrifugation speed (rpm)

36

synthetic wastewater itself in absence of adsorbent do not affect the dissolved amount of Ag-

NPs; hence 100% wastewater was selected for further adsorption experiments.

Figure 4.3 Effect of the strength of synthetic wastewater in solution on the

concentration of Ag in the water phase after a 24 hours mixing experiment.

4.3 Adsorption of Ag-NPs to different wetland substrates

Adsorption experiments with different wetland substrates were performed at two NPs

concentrations, 100 µg/L and 1000 µg/L. Figure 4.4 shows the amount of Ag adsorbed to each

substrate as a function of the initial nanoparticle concentration. As expected, at higher initial

concentration of Ag-NPs in the water phase all substrates adsorb a higher amount of Ag. At

initial concentrations of 100 µg Ag-NPs/l, the differences between the 3 substrates tested are

significant (p<0.05). Sand has the best retention capacity, with a total Ag amount of 0.25 ± 0.02

µg/g substrate being adsorbed. At higher initial concentrations (1000 µg/l Ag-NPs), differences

between adsorption to zeolites and sand (1.69 ± 0.14 µg/g substrate and 1.25 ± 0.02 µg/g

substrate , respectively) are not significant anymore (p=0.305), whereas adsorption to gravel

(0.10 ± 0.16 µg/g) differs significantly from adsorption to sand and zeolites (p<0.05). At all

concentrations, gravel displayed a lower adsorption compared to zeolite and sand.

0

5

10

15

20

25

0% 10% 50% 100%

Dif

fere

nce

in A

g co

nce

ntr

atio

n

(u

g/L)

Percentage of synthetic wastewater in solution

37

When examining adsorption to each substrate at different concentrations it can be concluded

that in the case of sand and zeolite significant differences can be found between 1000 µg Ag-

NPs/l and 100 and between 1000 µg Ag-NPs/l and 0 µg Ag-NPs/l (p<0.05), but no significant

differences are observed between the initial concentrations of 0 and 100 µg Ag-NPs/l (p=0.602

for sand and p=0.106 for zeolite). In the case of gravel, there were no significant differences in

the amounts of silver adsorbed between all concentrations tested (p=0.092).

Figure 4.4 Amounts of Ag adsorbed to sand, gravel and zeolite after a 24 hours, at

different initial concentrations of Ag-NPs.

4.4 Influence of biofilm on removal of Ag-NPs from wastewater

The effect of biofilm on the removal of Ag-NPs from wastewater was studied. Three different

gravel samples (with wet/active, dry/passive and no biofilm), were evaluated in the adsorption

experiments.

0.25

± 0

.02

0.0

6 ±

0.0

3

0.16

± 0

.02

1.25

± 0

.53

0.10

± 0

.16

1.69

± 0

.14

0

0.2

0.4

0.6

0.8

1

1.2

1.4

1.6

1.8

2

Sand Gravel Zeolite

Am

ou

nt

adso

rbed

ug

Ag/

g su

bst

rate

Substrate type

0 ug/l

100 ug/l

1000 ug/l

38

Figure 4.5 Effect of biofilm condition on silver adsorbed per gram of substrate.

As can be seen from the graph in Figure 4.5, the sample with wet/active biofilm presented the

highest adsorption and/or uptake (p<0.05), whereas presence of passive and no biofilm

resulted in lower and similar adsorbed amounts of silver, without any significant difference

(p=1.000). Figure 4.6 shows the amount of silver adsorbed per gram of biofilm. The total mass

of biofilm was calculated via loss by ignition (0.0508 g active biofilm/30 g gravel and

0.0365 g passive biofilm/30 g gravel). The amount of silver adsorption and/or uptake by the

biofilm was more than 2 times higher for the wet/active biofilm compared to the dry/passive

biofilm (p<0.05).

0.21 ± 0.03

0.06 ± 0.01 0.06 ± 0.03

0

0.05

0.1

0.15

0.2

0.25

Active Biofilm Passive Biofilm No biofilm

Am

ou

nt

ad

sorb

ed

ug

Ag/

g su

bst

rate

39

Figure 4.6 Effect of biofilm condition present in substrate on silver adsorbed per gram of

biofilm.

4.5 Adsorption to plant material and uptake by the plants

A 4-weeks adsorption and uptake experiment was performed in order to identify the different

processes that Ag-NPs may undergo in the presence of wetland plants. Table 4.3 shows the

amount of silver present in each phase (water, plant, glass recipient) and the total amount that

enters and leaves the system during this period, both for active and passive wetland plants.

Active wetland plants were selected as fresh, green looking plants from a pilot system,

whereas passive wetland plants were dry, dying plants from the same pilot. From the table it

can be seen that both for active and passive wetland plants, the mass balance was closed. The

amount of Ag entering the system was perfectly recovered in the amount of silver leaving the

system through the different phases (~100%). The total mass that enters the system

correspond to the sum of the amount of silver added each week. In the case of new and old

leaves, composite samples of the three replicates were analyzed, because of the low mass of

the material, and hence, no standard deviation (SD) can be shown. Also, passive wetland

plants experiments were performed using just the roots of the plant, and no old aboveground

tissue was present and analyzed. Nevertheless, contrary to what was expected this type of

121.1 ± 15.2

49.1 ± 10.3

0

20

40

60

80

100

120

140

160

Active Biofilm Passive Biofilm

Am

ou

nt

adso

rbed

ug

Ag/

g b

iofi

lm

40

plant were not completely passive and were able to grown during the 4 weeks period, hence

new aboveground tissue was analyzed.

Table 4.3 Mass balance of Ag-NPs during a 4-week experiment focused on adsorption to wetland

plants and uptake by the plants. Amounts of silver in- and out-of the system, considering all the

phases involved (average ± SD, n=3).

Mass balance Active wetland plant Passive wetland plant

µg Ag µg Ag

IN Total 73.09 ± 6.42 67.22 ± 0.90

OUT

Water phase 42.99 ± 11.23 28.13 ± 10.28

Adsorbed to glass 2.36 ± 0.52 3.08 ± 0.91

Adsorbed to roots 13.71 ± 1.62 27.63 ± 7.38

Uptake by roots 13.76 ± 8.15 8.75 ± 6.55

Translocated to new leaves 0.01 0.08

Translocated to old leaves 0.86 -

Total 73.12 ± 11.78 67.63 ± 11.35

Figure 4.7 shows the distribution of silver after the 4 weeks, both for active and passive

wetland plants. As can be seen from the graph the majority of the silver stays in the water

phase, however in the case of active wetland plants this fraction is remarkably higher. The

amount of silver adsorbed to the glass walls of the Erlenmeyer flasks containing the plants and

wastewater solution, was similar in both cases. However a striking difference can be seen for

the adsorption and uptake by roots. A higher adsorption to the roots was found in passive

wetland plants, whereas the uptake of silver by roots was significantly higher in active wetland

plants. The translocation of silver to aboveground tissue (new or old leaves) is negligible in

both cases, accounting for less than 1% of the total amount of silver.

41

Figure 4.7 Distribution of Ag-NPs between different phases after exposing the

NPs for 4 weeks to wetland plants; top: Active wetland plant; Bottom: Passive

wetland plant.

Water 58%

Adsorbed to glass 3%

Adsorbed to roots 19%

Uptake by roots 19%

Translocated to new leaves

0%

Translocated to old leaves

1%

Water 42%

Adsorbed to glass 4%

Adsorbed to roots 41%

Uptake by roots 13%

Translocated to new leaves

0%

42

5. Discussion

5.1 Effect of centrifugation speed and wastewater strength

Experiments were performed to evaluate the effect of centrifugation speed and wastewater

strength on the recovery of silver in the adsorption experiments. These results give an insight

in the aggregation and precipitation of silver compounds during the 24-hour period.

Physicochemical processes are assumed to be important in the removal of Ag-NPs in CW, since

they also account for the majority of the removed mass in the case of heavy metal ions (Lesage

2006). The experiments did not reveal significant differences in silver recovery after

centrifugation of the solution at 0, 500 and 1000 rpm after 24 hours constant mixing in Milli-Q

water. These results suggest that the loss of Ag-NPs due to aggregation, and subsequent

sedimentation of these aggregates is not an important process in Milli-Q water during the 24 h

in which the experiment took place. Also, it is important to take into account that the NPs used

for the experiments are coated with citrate, which prevents the autoaggregation of NPs

(Levard & Hotze 2012; Christian et al. 2008).

On the other hand, the results showed that the concentration of the used synthetic

wastewater did not have an effect on the recovery of silver after 24 h. It has been previously

demonstrated that Ag-NPs, after dissolution to silver ions, react with anions such as chloride

and sulfide to form insoluble compounds that precipitate from the aqueous phase (Levard &

Hotze 2012). Chloride is present in the synthetic wastewater used in the experiments. Sulfide

is not present, but sulfate may be transformed into sulfide under the anoxic conditions

occurring in CW (Lowry et al. 2012). However, the experiments were performed in oxic

conditions, decreasing the possibility of sulfide precipitation. The results obtained suggest that

the precipitation of silver salts is not an important process after 24 hours of reaction in the

synthetic wastewater used in this experiment.

43

5.2 Adsorption of Ag-NPs to different wetland substrates

The adsorption tests with different wetland substrates showed that sand and zeolite were able

to adsorb a higher amount of Ag-NPs than gravel. This can be due to the higher porosity and

CEC of zeolites, and the smaller size of sand particles in comparison to gravel (Moreno-Castilla

2004). Higher porosity, CEC and lower particle size increase the adsorption of pollutants, due

to a larger number of adsorption sites available and a larger surface to volume ratio. Also,

previous studies have shown a positive correlation between the organic carbon content and

the sorption potential of the sorbate (Brar et al. 2010). This was also observed in our study,

where sand and zeolites presented a higher content of organic matter than gravel. Ebeling et

al. (2013) have previously reported results of adsorption experiments for Ag-NPs in other

wetland substrates: silt loam soil, wetland soil and water treatment residuals (WTR). In

comparison, the silt loam soil retained less Ag-NPs than wetland soil and WTR. This can also be

attributed to the higher carbon content of wetland soil and WTR than silt loam soil. Ebeling et

al. (2013) also showed that adsorbed Ag-NPs were associated with the sand sized particles in

WTR, indicating that a higher removal can be reached with smaller sized particles.

Some studies on transport and retention of Ag-NPs in soil have previously been conducted.

Liang et al. (2013) have performed column experiments with loamy sand soil and surfactants

stabilized Ag-NPs, and have found that the retention of Ag-NPs was enhanced with increasing

ionic strength, decreasing flow rate and initial concentrations. Sagee et al. (2012) have also

concluded that the retention of Ag-NPs in the soil increases when the size of soil aggregates

increases and flow rates decrease. In the presence of nitrate instead of chloride, a decrease in

the retention of Ag-NPs was observed; hence the chemical interactions of Ag-NPs with soil are

of utmost importance to increase the removal of Ag-NPs.

When the effect of the concentration of the Ag-NPs was studied, a direct relation between the

initial concentration and the adsorbed concentration was seen for all substrates. The amount

44

of adsorbed silver was higher at higher initial concentrations of Ag-NPs. However, the

percentage of the initial silver concentration that was adsorbed was different between the

substrates. Sand and gravel show a decrease of the percentage adsorbed with increasing

concentrations (84.9% and 64.3% for sand, and 23.2% and 6.2% for gravel, at 100 µg Ag-NPs/l

and 1000 µg Ag-NPs/l, respectively). However, zeolite presents an opposite behavior, with

54.7% and 58.0% of the initial concentration of Ag-NPs adsorbed on the surface of the

substrate. In order to identify the maximum amount that each substrate is able to adsorb onto

its surface, it is necessary to perform adsorption tests during longer periods of time and a

wider range of concentrations. Adsorption isotherms are vital to identify the best sorbate and

the phenomena involved in the adsorption mechanisms (Moreno-Castilla 2004). The

evaluation of the adsorption and desorption processes in wetland substrates at a longer term

is essential to assess the efficiency of the removal of Ag-NPs from wastewater streams.

The different types of surface coatings that can be used to stabilize Ag-NPs also affect the

retention of Ag-NPs in soil systems (Lin et al. 2012). Even though surface coating changes the

surface structure and composition of NPs, it does not prevent chemical reaction, affecting just

the kinetics rates (Levard & Reinsch 2011). The surface coating on Ag-NPs studied by Lin et al.

(2012) included PVP and gum arabic. Even though the coating stabilizes nanoparticles against

aggregation, it enhances the retention by adsorption to a silica porous medium. Although in

the present work the effect of coating was not studied, it is important to consider when

evaluating the removal of Ag-NPs in constructed wetlands.

Both the results of this work and previous studies (Ebeling et al. 2013) performed on wetland

substrates and soil particles have proven that the retention of Ag-NPs is possible in wetland

systems by adsorption processes occurring during the water treatment. This means that CWTS

have the ability to remove Ag-NPs from wastewater. However, it is important to note that the

effect of aggregation and precipitation of Ag-NPs has not been accounted in this experiment;

45

hence the values of adsorption obtained may be overestimated. The composition of

wastewater will directly affect not just the adsorption of Ag-NPs on solid substrates, but also

the chemical processes that may take place, e.g. anions concentrations could affect

precipitation processes, and ionic strength could affect aggregation of stabilized Ag-NPs.

Further studies are needed to investigate the strength and kinetics of the adsorption of Ag-NPs

on wetland substrates.

5.3 Influence of biofilm on adsorption / uptake by micro-organisms

The effect of the presence of biofilm on the substrate was studied. A strong difference was

seen between the adsorption of Ag-NPs on wet/active biofilm versus no biofilm, with

0.21±0.03 µg Ag/g substrate and 0.06±0.03 µg Ag/g substrate, respectively. Sheng & Liu (2011)

performed experiments where biofilms were incubated in Ag-NPs suspensions. An evident

decrease in the concentration of Ag in the suspension was observed during the first

45 minutes, and a steady state was achieved after longer periods of time, indicating a

saturation of Ag-NPs in the biofilm after 520 minutes. As was the case for adsorption on solid

substrates, it is necessary to assess the interactions between biofilm and Ag-NPs during a

longer time to identify possible desorption, maximum adsorption capacity and possible

adverse effects of Ag-NPs on biofilm systems. When biofilms are exposed to heavy metals,

these can be easily immobilized by biosorption and chemical reactions with anionic ligands.

This is also the case for silver ions and Ag-NPs which are strongly adsorbed by bacterial cells

(Choi et al. 2010). Observations made by Choi et al. (2010) show that nanosilver is relatively

well distributed in a thin biofilm, although aggregates of Ag-NPs were detected, indicating that

the interactions of biofilm cells and Ag-NPs resulted in a strong aggregation.

Even though the fate of metallic NPs in wetland systems involving biofilm has not been studied

extensively before, Kaegi et al. (2013) have performed studies concerning the fate of Ag-NPs in

46

urban WWT systems. Electron microscopy images have proven that Ag-NPs do not aggregate

with each other in wastewater systems (i.e. no autoaggregation occurs). However, they

aggregate with organics present in the medium (heteroaggregation). Due to the high dilution

of Ag-NPs in the system, it becomes more likely that they will collide with other components

present (e.g. suspended solids, activated sludge particles). This implies that the Ag-NPs are

predominately attached to wastewater flocs, thus being effectively removed from wastewater,

independent of the treatment (aerobic, anoxic, anaerobic), size, coating (citrate or PVP) and

type of metallic nanoparticle (Ag versus Au). Hou et al. (2012) have also demonstrated that no

significant differences exist between the aqueous phase of control reactors without Ag-NPs

and reactors containing them. These results suggest that biosolids retain the majority of silver

that enters the system and are consistent with those found by Benn & Westerhoff (2008), who

concluded that biosolids in sludge will remove over 99% of the Ag-NPs present in wastewater.

In the case of urban wastewater systems, Ag-NPs are preferentially attached to organic matter

present in suspension than to the biofilm formed in sewer system (Kaegi et al. 2013). This can

be explained by the larger surface area provided by the suspended particles compared with

the biofilm. However, for SSF-CW this may not be the case. In this system a large amount of

solid substrate, such as gravel, is present. The large amount of biofilm that can be formed on

the surface of it will compete more effectively with the organics present in wastewater,

increasing the amount of Ag-NPs that can the adsorbed to the biofilm, compared to urban

wastewater systems. Moreover, CW operate at a long retention time, which can increase the

amount of adsorbed Ag-NPs due to a longer contact time between the water and solid phase.

In this experiment, the retention of Ag-NPs on wet/active and dry/passive biofilm was also

compared. A significant difference was found between wet/active and dry/passive biofilm

(0.21±0.03 µg Ag/g substrate and 0.06±0.01 µg Ag/g substrate), but no significant differences

are observed between dry/passive biofilm and no biofilm. Both samples, active and passive

47

biofilm, were taken simultaneously from the same pilot system, after which the passive sample

was dried at 40°C during 24 hours and the active sample was kept moist with synthetic

wastewater until the analysis. The application of heat on the previously formed biofilm may

produce changes in the cell wall (Özer & Özer 2003), which can affect the surface of cell

material and may affect the metabolic reactions in the cell. Besides adsorption of Ag-NPs to

biofilm, also the uptake of Ag-NPs by microbial cells can take place in the active cells. This

could explain the great difference encountered between the active and passive biofilm

(121.1±15.2 µg Ag/g active biofilm and 49.1±10.3 µg Ag/g passive biofilm). It is possible that

passive biofilm has been fully inactivated during the heating pre-treatment, considering that

microorganisms have an optimal temperature for survival and growth. Therefore, passive

biofilm would not be able to take up Ag-NPs via metabolic processes. However, this cannot

explain the difference encountered between the passive and the no biofilm scenarios

(0.06±0.01 µg Ag/g substrate and 0.06±0.03 µg Ag/g substrate for passive and no biofilm,

respectively). These results suggest that the activity of the biofilm played a significant role in

the retention of silver.

An aspect that has not been considered in this study is the possible negative effects that

Ag-NPs may exert on biofilm microorganisms in the long term, due to their bactericidal

characteristics. Potential adverse effects of Ag-NPs on the efficiency of removal of pollutants

and nutrients in urban wastewater treatment systems have been previously studied by

different research groups. Hou et al. (2012) demonstrated that the release of citrate coated

Ag-NPs into sewage would unlikely cause any adverse effects on the removal of COD and NH4

during activated sludge processes. The vulnerability of wastewater biofilm and planktonic

cultures to Ag-NPs was assessed by Sheng & Liu (2011) by exposing them to different doses of

Ag-NPs and exposure times. The results showed that wastewater biofilms are highly tolerant to

Ag-NPs; the reduction of heterotrophic plate counts was insignificant even after 24 hours of

exposure to 200 mg Ag/l. On the other hand, when treated as a planktonic culture, the same

48

isolated bacteria are highly sensible, where most bacteria died after 1 hour exposure to

1 mg Ag/l. These outcomes indicate that microbial community interactions play an important

role in the control of the bactericidal effect of Ag-NPs on wastewater biofilm. Another process

that decreases the toxicity of Ag-NPs in wastewater systems is the sulfidation of Ag-NPs. The

sulfidation of Ag-NPs creates a passivating layer that reduces the dissolution of them, hence,

decreasing the toxicity (Reinsch et al. 2012).

5.4 Plant adsorption and uptake

Plants in constructed wetlands provide surface for biofilm attachment and also for adsorption

and uptake of pollutants (Davis n.d.). During the four weeks that the experiment was

conducted, different transfer processes between plants and the water phase were studied,

both in active and passive plants. The mass balances show a good mass closure, however a

great deviation on the amount of Ag present in each phase (water, roots, leaves and glass) was

found. In all setups, the amount of silver adsorbed to the glass walls of the containers and

biofilm formed in it was negligible (3-4% of the total amount added).

In both cases, active and passive plants, the majority of the Ag-NPs were found in the water

phase after analyzing and balancing the amount of silver in the four weeks period. However, if

the amount of silver in the water phase is analyzed every week a different behavior can be

seen. In the case of active plants, a great adsorption and/or uptake to plant tissues can be

deduced, due to a small fraction remaining in the water phase; however after week 3 the

fraction in the water phase was over 100% of the initial amount added in the same period. This

behavior suggest that desorption of Ag-NPs previously adsorbed occurred, making wetland

plants both a sink and source of Ag-NPs. It is also possible that the plants have reached their

maximum level of accumulation, contributing to the increase of silver in the water phase. The

great standard deviation on the results is mainly caused by the difference in plant mass

49

present in the different replicates of the experiments. In order to understand more fully the

interactions between the plants and Ag-NPs present in the water phase, it is necessary to

perform experiments with longer times of exposure, and evaluate if plants can immobilize Ag-

NPs in their tissues efficiently.

The adsorption on roots was found to be an important process in the experimental systems,

accounting for the largest amount of sequestered silver, after the amount remaining in the

water phase. Roots provide a large surface area for attachment of pollutants, and for the

formation of biofilm that can also adsorb Ag-NPs on their surface (Davis n.d.). The amount of

Ag-NPs taken up by roots was the same as the amount adsorbed by roots in the case of active

wetland plants. However, for passive wetland plants a strong difference was found between

both processes. At the beginning of the experiment passive plants were considered as

dry/dying plants that would unlikely grow and take up a considerable amount of silver.

However, besides one of the triplicates, passive wetland plants grew just as active plants.

The amounts of Ag-NPs adsorbed by the roots showed a striking difference between active and

passive wetland plants (19% and 41%, respectively). On the other hand, the uptake of Ag-NPs

was higher in the case of active plants. It is possible that active plants were able to take up a

higher amount of silver, due to their stronger activity at the moment that they were taken out

from the pilot system and placed in the experimental set-ups in contact with Ag-NPs. The

adsorption and uptake of Ag-NPs by roots could have been hindered by a lack of water

movement during the experiment, which decreases the diffusion of silver to the roots, hence

decreasing the likelihood that roots will adsorb and take up silver from the surrounding phase.

In general, the amount of metal that can be accumulated in helophytes tissues has been found

negligible for the removal of metals in constructed wetlands (Lesage 2006). Harris & Bali

(2007) have studied the uptake of silver by two metal tolerant plants and have concluded that

the uptake of silver is directly dependent on the time of exposure and substrate metal

50

concentration, i.e. at higher concentrations and exposure times, the amount of silver taken up

increased too. This coincides with the results found by Xiao (2013), that studied the behavior

of cerium and silver nanoparticles with wetland macrophytes. The difference in uptake

between free ions and NPs was also assessed, concluding that plant exposed to free ions

accumulate more metals in their tissues. This is possible due to the smaller diameter of ions

compared with NPs, facilitating the transport through cell walls and membranes (Xiao 2013). It

is important to consider that the values found for the amount of silver taken up by roots, may

be overestimated as part of the amount adsorbed in the surface of them, could still be present

when the destruction and analysis of the roots was carried out due to possibly incomplete

detachment. The use of microscopic methods can be beneficial to identify whether silver has

been taken up or just adsorbed on the surface. However they were not available during this

study.

The amount of silver translocated to aboveground tissues is negligible in both cases, active and

passive plants, accounting for less than 1% of the total silver added to the system. This

behavior coincides with previously reported studies that suggest that Phragmites australis

restricts the translocation of metal to aboveground tissues (Weis & Weis 2004). Hence, most

wetland plants show higher metal concentrations in belowground biomass compared to

aboveground. The aboveground accumulation of metals generally account for less than 1% of

the metal removal, tested in different wetland plants (C. acutiformis, I. pseudacorus, J. effuses

and P. australis) (Lesage 2006).

5.5 General assessment of the removal processes in constructed wetlands

A number of processes interfere in the removal of Ag-NPs from CW systems. These processes

include sorption to plant roots, solid substrates and biofilm; uptake by plants and

microorganisms present; precipitation of dissolved silver as chloride and sulfide compounds

51

(AgCl and Ag2S); and aggregation and subsequent sedimentation of Ag-NPs (Sharif et al. 2013).

Most of these processes have been studied individually in the current research, except for

precipitation and aggregation, both of which represent a great fraction of the Ag-NPs removed

in this type of systems (Lowry et al. 2012).

Even though the results obtained in this research give an insight in the fate of Ag-NPs in

constructed wetlands for wastewater treatment, further studies are needed in order to

evaluate the efficiency of these types of systems in the removal of Ag-NPs. Further research

should be focused on the evaluation of the different processes on a long-term basis. For

example, desorption processes may take place from the solid substrate, biofilm or plant roots.

Also, Ag-NPs may exert toxic effects in the long term to microorganisms constituting the

biofilm and plants. Moreover, it is important to carry out an evaluation of the complete

system, i.e. study the fate of Ag-NPs when all processes are occurring at the same time, in

order to estimate more accurately the efficiency of the system for removal of Ag-NPs. Another

important aspect is the effect of different surface coatings on the removal of Ag-NPs in

constructed wetlands.

52

6. Conclusions and Recommendations

The general objective of this thesis was to evaluate the different processes that contribute in

the removal of Ag-NPs from wastewater in constructed wetlands. From the results found, it

was possible to conclude that the following:

1. The removal of Ag-NPs by adsorption to solid substrates is possible. Sand and zeolites have

shown to adsorb a higher amount of Ag-NPs than gravel, at the analyzed concentrations. The

adsorption of Ag-NPs on solid substrates depends on the organic matter content and size of

the substrate.

2. The biofilm formed on the surface of wetland substrates and plant roots contributes to the

removal of Ag-NPs by processes of adsorption and uptake. The presence of biofilm in wetland

substrates enhanced the removal of Ag-NPs from the water phase. A significant difference was

found between the adsorption/uptake on active and passive biofilm (0.21±0.03 µg Ag/g

substrate versus 0.06±0.01 µg Ag/g substrate), suggesting that the inactivation of

microorganisms has an important effect on the removal of Ag-NPs. However, no significant

differences were found between passive biofilm and experiments in absence of biofilm

(0.06±0.01 µg Ag/g substrate versus 0.06±0.03 µg Ag/g substrate), suggesting that the uptake

mechanisms account for most of the removal of Ag-NPs from the water phase.

3. The majority of Ag-NPs remained in the water phase in presence of plants after a period of 4

weeks. However, until week 2 an important amount of silver was removed from the water

phase via adsorption and/or uptake by the roots. Evidence of desorption and/or maximum

accumulation of silver on plant roots was suggested by an increase of Ag-NPs after the third

week, especially in active wetland plants. The adsorption of Ag-NPs by roots accounted for a

greater percentage of silver mass than uptake by roots, and the translocation to aboveground

53

tissues is negligible in wetland plants (<1%). Active wetland plants show a higher uptake of Ag-

NPs and lower adsorption on roots compared to passive wetland plants.

Even though the present research contributes to the understanding of the different processes

that participate in the removal of Ag-NPs in constructed wetlands, further studies are needed

to fully comprehend the fate of Ag-NPs in constructed wetlands for water treatment. The

efficiency of removal will depend on the different processes individually and combined, hence

experiments that combine the different processes should be performed to evaluate the

performance of constructed wetlands in the removal of Ag-NPs. It is also important to evaluate

the different processes during a longer period of time in order to identify the maximum

adsorption capacity of the substrates, biofilm and plant roots and possible desorption. In the

current study, aggregation and precipitation processes were not considered; however, they

are vital in the removal of Ag-NPs from wastewater systems, and it is necessary to evaluate the

extent of these reactions in constructed wetland systems.

54

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