landfill leachate treatment: a new photobioreactor technology mct… · table 2 - advantages and...

Landfill leachate treatment: a new photobioreactor

technology

João Alexandre Bastos Sousa

Dissertação de Mestrado em Contaminação e Toxicologia Ambientais

2010

João Alexandre Bastos Sousa

Landfill leachate treatment: a new photobioreactor technology

Dissertação de Candidatura ao grau de Mestre em Contaminação e Toxicologia Ambientais submetida ao Instituto de Ciências Biomédicas de Abel Salazar da Universidade do Porto.

Orientador – Prof. Doutora Olga Maria Lage

Categoria – Professora Auxiliar

Afiliação – Faculdade de Ciências da Universidade do Porto

Co-orientador – Dr. Nuno Gomes

Categoria – Administrador/Investigador

Afiliação – Bluemater S.A.

João

Sousa

Porto 2010

Lan

dfi

ll lea

ch

ate

tre

atm

en

t: a

new

ph

oto

bio

reacto

r te

ch

no

log

y

Acknoledgements

I would like to thank all those who supported my work and contributed to the best

developing of it:

To Prof. Doutora Olga Lage for all the unconditional support and encouragement

that made all this work possible regardless of the many setbacks occurred, and also for

allowing me to explore new ideas which contributed greatly to my motivation.

To Dr. Nuno Gomes for the opportunity to work with this thrilling new technologies

and for all the support and commitment towards the study.

To Dr. Newton Gomes for all the assistance in the thesis planning and opportunity

to learn new methods.

To Dr. Joana Bondoso for all the help and tips in the laboratorial work.

To the FCUP, LEMAM and Bluemater S.A. colleagues that contributed to the

accomplishment of this thesis.

And finally, a special thanks to my parents that always supported me and never

backed down in any situation.

Abstract

The landfill leachates represent a major environmental problem and their

treatments present big challenges. Many technologies have been developed but the quest

for the best treatment technology is still ongoing. The microalgae treatment technology

has been gaining momentum due to their advantages. In this thesis, a pilot treatment plant

with an innovative configuration and a new photobioreactor of attached biomass was

tested and its startup monitored. Three photobioreactor diameters were tested. The

assembling process was phased and diverse operational problems occurred, leading to

complex analysis of the results obtained. The COD, BOD, total N, NH4-N, NO2-N, NO3-N,

PO4-P, Fe, DO, pH and temperature have been analyzed in the various treatment

components to access the treatment efficiency of each component and of the whole

system. The overall efficiency proved to be low, attaining the COD and the NH4-N mean

removal rates of 6.4% and 19.1% respectively, and no difference was observed between

the different photobioreactors. The biofilm formation inside the photobioreactors was

studied by optical and microbiological methods and results showed that complex

dynamics of the microbial communities occurred. The dominant microalga present in the

photobioreactors, Chlorella sp., was isolated for further characterization.

Resumo

Os lixiviados de aterros sanitários representam um importante problema ambiental

e o seu tratamento continua a ser um grande desafio. Até à data várias tecnologias têm

sido desenvolvidas. Contudo a procura pela tecnologia mais eficaz permanece um

problema em aberto. Devido às vantagens apresentadas, as tecnologias de tratamento

usando microalgas têm vindo a ganhar destaque. Assim, neste trabalho de tese foi

contruída e testada uma estação piloto de tratamento de lixiviados com uma configuração

inovadora e com um novo tipo de fotobiorreactor. Procedeu-se à monitorização do seu

funcionamento desde a fase de arranque. Foram testados fotobiorreactores com três

diâmetros diferentes. O processo de montagem foi faseado. Diversos problemas

ocorreram ao longo deste processo, levando a uma análise complexa dos resultados

obtidos. Os seguintes parametros, COD, BOD, N total, NH4-N, NO2-N, NO3-N, PO4-P, Fe,

DO, pH e temperatura, foram analisados nos vários componentes do sistema para avaliar

a eficiência do tratamento em cada componente e no seu todo. A eficiência global

demonstrou ser baixa. Taxas médias de remoção de 6.4% e 19.1% foram obtidas

respectivamente para os parâmetros COD e NH4-N. Não foram observadas diferenças

entre os diferentes tipos de fotobiorreactor. A formação do biofilme nos fotobiorreactores

foi estudada por métodos ópticos e microbiológicos tendo os resultados mostrado existir

uma dinâmica complexa na comunidade microbiana. A microalga dominante presente no

fotobiorreactor, Chlorella sp., foi isolada para posterior caracterização.

Index

Figures index……………………………………………………………………………………….1

Tables index………………………………………………………………………………………..4

List of abbreviations……………………………………………………………………………….5

Introduction…………………………………………………………………………………………6

Landfill leachate problem………………………………………………………………...7

Leachate treatment technologies……………………………………………………….8

Physic-chemical treatments……………………………………………………………...9

Biological treatments …………………………………………………………………...10

Microalgae for wastewater treatment………………………………………………….13

The algae-bacteria method……………………………………………………………..14

Advantages of this method to the landfill leachate treatment………………………16

Photobioreactors for wastewater treatment…………………………………………..16

The algae-bacteria community relevance…………………………………………….17

Objectives………………………………………………………………………………...19

Chapter I - Pilot leachate treatment plant planning and assembling……………………….20

Leachate chemical characterization…………………………………………………..21

Problems of biological treatment………………………………………………………22

Pilot treatment plant planning………………………………………………………….22

Assemblage………………………………………………………………………………28

Circuit description……………………………………………………………………….28

Operational data………………………………………………………………………...30

System operational problems………………………………………………………….31

Chapter II - Physic-chemical monitoring of the pilot leachate treatment plant startup……32

Material and methods…………………………………………………………………...33

Results and discussion…………………………………………………………………34

o Initial remarks……………………………………………………………………34

o Untreated leachate……………………………………………………………..35

o Ozone treatment………………………………………………………………..36

o Trickling filter treatment………………………………………………………...37

o Clarifier treatment……………………………………………………………….39

o PBRs treatment…………………………………………………………………40

o Pilot treatment plant results……………………………………………………43

o Treatment comparison between first and second O3 generators…………46

Chapter III - Photobioreactors biofilm characterization………………………………………48

Material and methods…………………………………………………………………..49

Results and discussion…………………………………………………………………51

o Initial notes………………………………………………………………………51

o In situ observations……………………………………………………………..51

o Sampling plates optical analysis………………………………………………53

o Microalgae and heterotrophic bacteria dynamics…………………………...57

o Microalgae isolation…………………………………………………………….59

Conclusions……………………………………………………………………………………….61

References………………………………………………………………………………………..64

1

Figures index

Figure 1 - Number of publications with the topic “landfill leachate treatment”.

Figure 2 - Microalgae and heterotrophic bacteria interactions.

Figure 3 - LIPOR I landfill, Ermesinde.

Figure 4 - Pilot treatment plant scheme.

Figure 5 - O3 treatment system scheme.

Figure 6 - A - Sycon® plate; B - Sycon® plates disposition.

Figure 7 - Trickling filter scheme.

Figure 8 - Laminar clarifier scheme.

Figure 9 - PBR scheme.

Figure 10 - PBRs substrates. a - polycarbonate alveolar plates; b - position of the

polycarbonate alveolar plates in the PBRs.

Figure 11 - PBRs disposition.

Figure 12 - Leachate treatment plant circuit scheme.

Figure 13 - Pilot leachate treatment plant.

Figure 14 - Pilot leachate treatment plant with the second O3 generator.

Figure 15 - Sampling locations.

Figure 16 - Untreated leachate (A) pH, temperature; (B) COD, total N, NH4-N; (C) PO4-P

and Fe variation.

2

Figure 17 - NO2-N and NO3-N concentrations before and after the O3 treatment and

removal percentage by this treatment.

Figure 18 - DO before and after the O3 treatment.

Figure 19 - COD, total N, NH3-N and PO4-P removal rates with the trickling filter

treatment.

Figure 20 - NO2-N and NO3-N concentrations before and after the trickling filter treatment

and production percentage by this treatment.

Figure 21 - COD, total N, NH4-N and PO4-P removal rates by the clarifier.

Figure 22 - NO2-N and NO3-N removal rates by the clarifier.

Figure 23 - COD, total N, NH4-N and PO4-P removal rates by the whole PBR system.

Figure 24 - (A) COD, (B) total N, (C) NH4-N and (D) PO4-P average removal rates from

the three replicates of the different PBRs.

Figure 25 - pH and temperature variation before and after the PBR treatment.

Figure 26 - COD, total N, NH4-N and PO4-P removal rates by the pilot treatment plant.

Figure 27 - Leachate (A) COD, (B) total N, (C) NH4-N, (D) NO2-N, (E) NO3-N, (F) PO4-P,

(G) Fe, (H) pH, (I) temperature and (J) DO evolution along the treatment plant

and time.

Figure 28 – Comparison of COD, total N and NH4-N removal percentages by the whole

treatment between the two O3 generators.

Figure 29 - Comparison of (A) NO2-N and (B) NO3-N production by the whole treatment

between the two O3 generators.

Figure 30 - Polycarbonate plates of 1 cm2 placed inside the PBR (a) before filling the PBR,

(b) after filling the PBR, and (c) 4 days after functioning.

3

Figure 31 - Schematic representation of the biofilm formed on the walls of the PBRs,

observed from the outside.

Figure 32 – Biofilm detachment in the PBRs: (A) small detachment spots and (B) big

detachment portions.

Figure 33 - Color images of the sampling plates.

Figure 34 - Evolution of the pigmented biofilm coverage on the sampling plates.

Figure 35 - Biofilm thickness profiles of the sampling plates.

Figure 36 - Black and white images of the sampling plates.

Figure 37 - Evolution of the mean gray values of the sampling plates.

Figure 38 - Evolution of the microalgae cells/cm2 during the experiment.

Figure 39 - Evolution of the heterotrophic bacteria during the experiment expressed in

CFU/cm2.

Figure 40 – Microalgae isolated from the phototrophic biofilm.

4

Tables index

Table 1 - Factors influencing algae growth

Table 2 - Advantages and disadvantages of the treatment systems for landfill leachate

treatment.

Table 3 - Volume and surface area available for biofilm growth in the PBRs.

Table 4 - Hydraulic retention times.

Table 5 - HACH-Lange kits used for chemical analysis.

Table 6 - Positioning of the sampling plates inside the PBR.

5

List of abbreviations

BOD – Biological oxygen demand

CFU – Colony forming unit

COD – Chemical oxygen demand

DO – Dissolved oxygen

EC50 – Effective concentration 50%

EEA – European Environmental Agency

EPA – Environmental Protection Agency

EPS – Exopolymeric substances

EU – European Union

HRT – Hydraulic retention time

HRTP – High rate treatment pond

LC50 – Lethal concentration 50%

M.O. – Optical microscope

PAH – Polycyclic aromatic hydrocarbon

PBR – Photobioreactor

Total N – Total nitrogen

Introduction

6

Introduction

Introduction

7

Introduction

Landfill leachate problem

The high amount of waste generated by the world population represents a major

environmental problem. In the EU, approximately 1.3 billion tonnes of waste was

produced in 2002 which correspond to approximately 2.9 tonnes per person (Eurostat

2005). The development of under-developed countries, the increase of the world

population and urbanization will consequently lead to an increase of waste disposal all

around the world. This can be shown by the existing correlation between the economic

growth and the waste generation (EEA 2001). Landfills still represent one of the main

endpoints of the population solid wastes. In EU, approximately 100 million tones of

municipal waste were sent to landfills in 2007. In US the number of landfills is decreasing

but on the contrary, the size of the new landfills is increasing (EPA 2009). Overall, landfill

still remains the more economically attractive solution for waste disposal. The main

drawbacks of landfills are the land usage, production of high amounts of methane and the

production of highly contaminated leachates. Except for the land usage, the other two

problems can be minimized. The methane produced can be consumed and used for

energy production and the leachate can be treated. Leachate treatment however

represents a big challenge due to its highly complex and variable composition. This

complexity poses a great difficulty to the employment of traditional treatment used for

other kinds of wastewater.

Within the complex composition of the landfill leachate, there are some

compounds characteristic that are usually present: ammonia-nitrogen, humic substances,

heavy metals and phenols (Baun et al. 2004, Renou et al. 2008). A vast number of

xenobiotic substances can also be found, e.g. pharmaceuticals, pesticides, plasticizers,

chlorinated aliphatics and aromatic compounds (Schwarzbauer et al. 2002, Baun et al.

2004). Even though these ones only account for approximately 1% of the total organic

carbon content (Baun et al. 2004), it should be noticed that many of these contaminants

may be harmful to the environment even in very low quantities.

The leachate composition varies depending on many factors, such as the type of

waste disposed in the landfill, the age of the landfill, the precipitation, the season of the

year and the construction characteristics. In particular, the landfill’s age has a significant

impact on the leachate composition. The young landfills usually produce leachates with

high amounts of ammonia-nitrogen and biodegradable organic matter, and some

refractory organic matter. The old landfill’s leachates usually contain less ammonia-

nitrogen and organic matter but the majority of this is non-biodegradable (Huo et al. 2008).

Introduction

8

This difference is usually marked by the transition from the acid phase to the

methanogenic phase. In the acid phase, the ammonia levels rise due to aminoacids

degradation and fermentation leads to an increase of volatile acids. During the

methanogenic phase, the organic matter is mostly humic and fluvic fractions (Kulikowska

and Klimiuk 2008) which are usually non-biodegradable.

Toxicity tests performed with various types of organisms confirm that leachates

represent a potential danger to the receiving water stream (Bernard et al. 1996). Acute

and chronic toxicity tests performed with Ceriodaphnia dubia resulted in a 24h EC50 of

25% leachate concentration and a 48h EC50 of 14% leachate concentration (Dave and

Nilsson 2005). It was also shown that the high ammonia levels were the main cause of

toxicity for C. dubia and, when the leachate is treated inefficiently, the high nitrite amounts

produced by an incomplete nitrification can result in an increase of toxicity (Dave and

Nilsson 2005). In an acute toxicity test with a fish model, Brachydanio rerio, the 48h LC50

and 96h LC50 was 2.24% of leachate concentration, confirming the high toxicity of the

leachate (Bila et al. 2005). Genotoxic and reproductive toxicity were also shown (Sang

and Li 2004, Dave and Nilsson 2005).

Leachate treatment technologies

In order to eliminate this danger to the environment, suitable treatments must be

employed to efficiently treat these complex leachates. Studies concerning landfill leachate

treatment have increased in the recent years (Figure 1).

Figure 1 - Number of publications with the topic “landfill leachate treatment”. Source: ISI

Web of Knowledge.

Introduction

9

The major challenge in treating landfill leachate consists in the high variability of

the leachate and high amounts of toxic compounds. A good characterization of the

leachate prior to planning the treatment is essential in order to choose the most suitable

treatment for each case. Biological treatment is the most common wastewater treatment

due to its advantages such as low cost and simplicity. However, it presents many

drawbacks for treating landfill leachates. The elevated concentrations of ammonia and

other toxic substances may inhibit the biodegradation. The low biodegradability of the

organic content also poses a difficulty to this method. In order to overpass the difficulties

of the biological treatment, a great development has been made in the physic-chemical

treatment process. A wide variety of technologies are currently available, but all of them

have some drawbacks associated.

Physic-chemical treatments

Coagulation-flocculation process has been successfully used to treat landfill

leachate (Amokrane et al. 1997). Nonetheless, this process has the disadvantage of

producing high amounts of toxic sludge and increase the concentration of metals in the

liquid phase (Silva et al. 2004). The cost of the coagulants and the sensitivity to pH are

also considered disadvantages of this process (Kurniawan et al. 2006).

Chemical precipitation has also showed good results, and has the advantage of

the low cost of the precipitants, the simplicity of the process and the value of the

precipitated when struvite is employed (Ozturk et al. 2003). However, the problems

mentioned for the coagulation, except the coagulant cost, are also applicable to this

treatment (Kurniawan et al. 2006).

To remove high amounts of ammonia, the ammonium stripping is a widely used

technique, but only has good results for ammonia removal. The chemical oxygen demand

removal is low and it is necessary to adjust pH which constitutes an extra cost (Kurniawan

et al. 2006).

The activated carbon adsorption has very high efficiencies but the need of

regeneration of the activated carbon has an elevated cost (Kurniawan et al. 2006).

Chemical oxidation has been used to treat refractory compounds in landfill

leachate. The majority use simple ozonation, but for higher efficiencies, combinations of

oxidants have been employed (Wu et al. 2004). Chemical oxidation is able to oxidize

organic substances to their highest stable oxidation states and improves the

biodegradability of recalcitrant organic matter (Renou et al. 2008). The major drawback of

this process is the high electrical energy demand, leading to an increase in the treatment

cost. In order to reduce the cost, this technique may be used with a complementary

Introduction

10

biological treatment. Nonetheless, the production of some oxidation products may

increase the leachate toxicity and inhibit the biological treatment (Lopez et al. 2004).

The electrochemical treatment is also used to treat leachates with good results

(Bashir et al. 2009). However the high cost of the treatment and the decomposition of

some key components make it a less viable treatment for landfill leachates (Renou et al.

2008).

Many technologies of membrane filtration have been improved recently.

Ultrafiltration, nanofiltration and reverse osmosis are some of the most used technologies

for landfill leachates (Renou et al. 2008). From all these, the reverse osmosis has the best

decontamination efficiency and seems the most promising technology for landfill leachate

(Linde et al. 1995). But there are negative aspects concerning the reverse osmosis

(Renou et al. 2008). The membrane fouling implicates an extensive pre-treatment or

chemical cleaning, and this can shorten their lifetime. Also, the filtration generates large

volumes of concentrated contaminants which are highly toxic waste. The issue with

filtration technology is that the contaminants are not degraded but concentrated, and the

waste generated has to be treated elsewhere.

Biological treatments

The biological treatments offer many advantages as low operational costs, the

production of more valuable sludge, and the possibility of using the vast metabolic

capacities of organisms for degradation of a wide range of contaminants. Whatever the

specific technology, the biological treatment consists in implementing measures to

optimize the growth and activity of the microorganisms that are responsible for the

wastewater treatment. Complex natural communities are used due to their stability and

adaptation to the wastewater composition. In some treatments, specific organisms may be

added or an inoculation with sludge from other locations may be applied to improve the

treatment capacity. As mentioned above, the use of biological treatment technologies in

landfill leachate poses a great challenge due to the leachate characteristics. However,

new developments in this field are turning the biological treatment processes more robust

and capable.

Within the biological treatment, there are many different technologies with their

pros and cons. The biological treatment of landfill leachate may be divided into three main

treatments: leachate transfer, aerobic treatment and anaerobic treatment.

The leachate transfer consists in moving the leachates to other treatment facilities,

like domestic sewage treatment plants, or recycling it back into the landfill where it will be

treated by the microbial community present. These two processes are some of the

simplest processes but have some disadvantages. The recycling method has been proven

Introduction

11

efficient but it may cause inhibition of the anaerobic metabolism, inhibiting biodegradation

and methane production (Ledakowicz and Kaczorek 2004), accumulation of volatile acids,

flooding and clogging in the landfill (San and Onay 2001). The leachate transfer to

domestic sewage plants may have prejudicial effects in the sewage treatment due to the

toxic compounds, refractory organic matter and high ammonium levels that may cause

inhibition of microbial metabolism. To avoid these problems, specific treatment plants

should be designed for treating the leachates, either aerobic or anaerobic.

In the aerobic treatment, the growth of aerobic organisms is promoted. These

microorganisms use the organic matter and dissolved oxygen present in the wastewater

for their metabolism and growth, thus removing organic contamination. Two main

configurations may be used for the aerobic treatment: the suspended biomass, such as

aerobic activated sludge or aerated lagoons, and the attached biomass, such as

membrane filters and trickling filters.

The most common wastewater treatment plants use aerobic treatment due to its

simplicity and efficiency, being the aerobic activated sludge the most used technique. The

activated sludge consists in the use of biological flocs containing inorganic matter, organic

matter and microorganisms that are responsible for the degradation of the contaminants

and include bacteria, fungi, protozoa and other living forms. In this process, the flocs are

circulated in order to be in contact with the organic material and oxygen. This implies a

great amount of energy to maintain the circulation and aeration of the system which

constitutes one major disadvantage. Other disadvantages are related to the fragility of the

organisms present in the process. If the optimum conditions of aeration are not

maintained properly, if a high variation in physic-chemical characteristics occurs or if the

wastewater has a punctual contamination with a highly toxic compound, the microbial

activity may be decreased, inhibited or the whole biological community may even

collapse. Due to the landfill leachate highly variable composition, high ammonium levels

and low biodegradability, this process is less suited.

Another aerobic treatment method is the aerated lagoons. These are easy to build,

easy to maintain and have good COD and ammonium removal efficiencies. However, the

use of aerated lagoons requires a big area available for its construction and problems as

odors, algae blooms or insect infestation may occur in this system (Wang 2009) making it

unpleasant for the surrounding population and environment.

Trickling filters make use of the biofilm growth on a fixed substrate used to treat

water that runs through it. There are many substrates and configurations available that try

to maximize the surface area available for biofilm formation and at the same time prevent

the clogging of the system. This biofilm has some advantages over the use of suspended

biomass. The attached biomass tends to be more resistant to toxic compounds, high

Introduction

12

ammonium content and low temperatures. This resistance also allows the growth of a

more stable microbial diversity, improving thus its metabolism and the degradation of

water pollution. The basic biofilm structure consists of an anaerobic layer present near the

substrate and an aerobic layer near the water-biofilm interface. This layer organization

occurs due to the limit of DO diffusion in the biofilm. The aerobic layer, usually the first to

colonize the substrate, is normally the responsible for the organic matter degradation and

the removal of ammonium and other contaminants. The first organisms to attach to the

substrate are responsible for the production of EPS that allow the attachment of other

organisms. When the thickness of the biofilm reaches the point of maximum dissolved

oxygen diffusion, the anaerobic layer starts to form near the substrate-biofilm interface.

This layer will decompose organic matter present in the biofilm and with gas production,

which will be responsible for the detachment of the biofilm when it reaches a certain

thickness. The detachment will open a space for new biofilm formation (Wang 2009). This

ecological succession creates a productive dynamics in the filters and allows the

degradation of biodegradable compounds and accumulation of non-biodegradable

compounds in sludge form. However the amount of sludge formed is reduced when

compared to suspended processes due to the biofilm increased stability. This constitutes

an advantage because there is less sludge to dispose of. The higher resistance of the

biological community makes this system more suitable for treating landfill leachate.

Previous studies showed a good efficiency even without complex operational conditions

which increase the costs (Matthews et al. 2009). More efficient substrates and

configurations could increase further the treatment efficiency turning this method even

more advantageous.

The anaerobic treatment consists on generating the optimum conditions for

anaerobic organisms to thrive. As in the aerobic treatment, there are technologies for

suspended biomass and attached biomass. The attached biomass offers the advantages

already mentioned above and has been shown to generate better removal rates (Renou et

al. 2008). The anaerobic treatment has the advantage of producing fewer solids, efficiently

removing BOD and generating methane which may be used for energy production. It is

more suited than aerobic treatment for treating high strength wastewaters (Chan et al.

2009). The lower needs of phosphorous and energy simplifies the operation of this

technology. On the other hand, it is known for its low reaction and increased fragility to

high ammonium concentrations, variations in pH and temperature and presence of heavy

metals (Wiszniowski et al. 2006). The low ammonium removal rates (Chan et al. 2009)

make this system not suitable for treating landfill leachates.

Introduction

13

Microalgae for wastewater treatment

Wastewater treatment using the capacities of microalgae is an old idea (Oswald et

al. 1955) that is gaining momentum nowadays due to several advantages in comparison

to other biological technologies. Microalgae are rapid growing photosynthetic organisms

that use sunlight as a source of energy and use nutrients such as nitrogen and

phosphorous to grow. Several microalgae may even act as heterotrophic organisms in

conditions where light is not available. These metabolic capacities make the microalgae

great candidates for wastewater treatment. They remove the nitrogen and phosphorous

by incorporating this nutrients in their biomass as they grow. The fact that microalgae use

the light as energy source rends these organisms suitable for removing nutrients when

organic carbon, a chemical energy source, is not available in sufficient amount. As

microalgae growth increases the water pH, some stripping phenomena may take place

contributing to additional nitrogen removal.

It is also known that microalgae accumulate heavy metals in its biomass, thus

removing them from the water, which poses a great advantage for treating water

contaminated with these toxic contaminants. pH and temperature increase of the water

due to photosynthesis of microalgae may also contribute to the elimination of pathogenic

bacteria (Fallowfield et al. 1996).

Adding to the removal capacities already mentioned, the microalgae systems offer

other operational advantages. As they consume CO2 and produce O2, they may be used

to oxygenate water with low DO, improving posterior heterotrophic biological treatment

with aerobic microorganisms and eliminate the need for aeration (Oswald 1988). This

eliminates the elevated costs of aeration which may represent 45-75% of the total energy

consumption of the treatment plant (Larsdotter 2006). The production of O2 and

consumption of CO2 makes this system carbon negative rendering it more environmentally

friendly by contributing to the reduction of the greenhouse gases, and gaining CO2 credits

in the CO2 emissions market. Microalgae treatment also eliminates the need of additional

treatments with other chemicals which leads to a reduction in sludge production. The

microalgae sludge has the advantage of being an energy and nutritionally rich sludge,

making it suitable for energy (Brennan and Owende 2010), fertilizer or feeding (Spolaore

et al. 2006, Mata et al. 2010) downstream applications. The possible use of this type of

sludge may reduce even further the operational costs of the microalgae system.

The microalgae are able to support adverse conditions such as strong variations in

pH, low temperatures and different salinities (Grönlund 2002, Oilgae 2009). They are easy

to control although some more knowledge is needed to understand its optimum

functioning. The microalgae growth is conditioned by several factors (Table 1) but some

are more resistant than others to these. The microalgae that already exist in the

Introduction

14

wastewater to be treated are usually the most suited for treating it (Jiménez-Pérez et al.

2004). This is due to the tolerance that these microalgae develop while thriving in these

stressful environments.

Table 1 - Factors influencing algae growth (Becker 1988)

Type Factor

Abiotic

Light

Temperature

Nutrient concentration

O2, CO2

pH

Salinity

Toxic chemicals

Biotic

Pathogens

Predation

Competition

Operational

Mixing

Dilution rate

Depth

Harvesting frequency

The microalgae may be used as suspended biomass or attached biomass. The

difference between these two systems has been explained before, and the same

principles apply for the microalgae. The attached form has improved resistance to

conditioning factors, reduces the sludge formation and in the algae case, the sludge

formed by attached biomass is much more easily settled than the suspended biomass

(Guzzon et al. 2008) which constitutes a great advantage of this system. Another

advantage of algal biofilms is that these accumulate suspended solids in contrast to

suspended algae that are difficult to remove and also contribute to an increase in

suspended solids and BOD in the final effluent. The big challenge with phototrophic

biofilms is the light penetration that has to be considered when choosing the technology to

use.

The algae-bacteria method

The microalgae used as a pure culture for wastewater treatment has many

problems due to contamination by other organisms. However it has been shown by

Introduction

15

several authors that the coexistence of microalgae and bacteria may provide even better

results than pure cultures (Borde et al. 2003). This symbiotic relation has been found to

produce good results for the removal of organic matter, ammonium, phosphorous (de-

Bashan et al. 2002) and other pollutants as PAHs, salicylate and phenols (Borde et al.

2003, Safonova et al. 2004, Chavan and Mukherji 2008).

The interaction between microalgae and bacteria is quite complex (Figure 2) and

few is known about it. Microalgae are responsible for the O2 production used by aerobic

bacteria to biodegrade organic pollutants with CO2 release that on its turn will be used by

microalgae to grow. Microalgae are also known to secrete EPS that may provide

conditions for heterotrophic bacteria to attach and serve also as food source (Muñoz and

Guieysse 2006). On the other hand, pH and temperature increase due to phototrophic

activity may have a negative impact on heterotrophic bacteria (Oswald 2003). Microalgae

and bacteria are known to secrete a wide range of compounds which may have effects on

the relation between them. Microalgae produce toxins that affect other organisms, like

bacteria, compromising their growth (Oswald 2003) but also secrete other metabolites that

enhance growth (Wolfaardt 1994). Concomitantly, bacteria also secrete compounds that

enhance microalgae growth and activity, as seen in studies using Azospirillum brasilense

(de-Bashan et al. 2004). However, bacteria may also secrete algaecides that inhibit

microalgae (Fukami 1997). These complex relations should be better understood in order

to take advantage of the enhancing effects and improve this system even further.

Figure 2 - Microalgae and heterotrophic bacteria interactions.

Introduction

16

Within this system, O2 production by microalgae is often considered the limiting

factor for contaminants removal, which is due to the slower growing rate of microalgae

compared to the growing rate of heterotrophic bacteria (Muñoz et al. 2004). Due to this,

high O2 production microalgae should be used to generate better results.

Advantages of this method to the landfill leachate treatment

This method has already been applied to landfill leachate with very good results

(Lin et al. 2007). Microalgae combined with bacteria form biofilms that are more resistant

to toxic compounds and high ammonium concentrations. The low biodegradability of the

organic matter present makes the microalgae very suitable for the treatment. However,

the few biodegradable organic matter has to be removed, which is accomplished by the

bacteria within this system. The inhibitory action of some pollutants may be minimized by

the symbiotic relationship which, as mentioned above, speeds up the biofilm development

due to algae growth promoters produced by bacteria.

Photobioreactors for wastewater treatment

The microalgae wastewater treatment is usually applied through the use of high

rate algae pond (Oswald 1988) which works as an open PBR. These ponds are artificial

ponds built to maximize microalgae growth and wastewater treatment. Their building and

operating low costs and the good efficiency make these the most used technique for

microalgae treatment. However there are some nuisances inherent to this system as in

the case of the aerated lagoons. In addition, it requires elevated hydraulic retention times

and short depth. These two factors increase the area needed for building an efficient high

rate algae pond which may turn this treatment option not suitable in some situations.

The other option to use microalgae is through other PBR configurations that may

give the same results using less area. This is the case of vertical PBRs that are usually

built with a closed configuration which have been shown to generate good organic matter

and nutrients removal rates (Molinuevo-Salces et al. 2010). The closed configuration

leads to a lower loss of water by evaporation, better control and higher efficiencies (Pulz

and Pulz 2001). The issue with this kind of PBRs is the elevated costs of building, difficulty

of operation and scale-up. However, major developments have been made in recent years

and these problems have been minimized. A wide variety of configurations are

commercially available for algae biomass production. However, there are less

configurations specialized for wastewater treatment. The objectives of the two activities

are similar but in the case of wastewater treatment, the suspended algae cells are

Introduction

17

considered a problem as already mentioned above. For this reason, PBRs for immobilized

or attached microalgae are more suited for this application.

The immobilized microalgae consist in immobilizing the microalgae cells in a matrix

which are then placed inside the PBR. This system eliminates the difficulties for

clarification and has been tested with many matrixes with good results (Moreno-Garrido

2008). However these tests were all performed at laboratory scale because immobilization

procedure is complicated and increases the treatment costs (Muñoz et al. 2009).

The attached microalgae PBR only requires surface available inside the PBR

which turns the system more simple and still eliminates the difficulties for clarification.

This type of PBR also eliminates the need for mechanical mixing because the mixing

created by the wastewater entrance is sufficient to dilute the contaminants. Nonetheless,

a gradient may be formed if the PBR is too high what will lead to a variation in the biofilm

communities present on different depths of the reactor. The surface/volume ratio and the

light penetration are very important factors when designing this type of PBRs. The more

surface is available, more biofilm will form and better efficiency may be obtained.

However, light penetration imposes a limit to the surface available in a PBR due to the

light needed by phototrophic biofilms to develop. The excess of biomass may also be an

obstacle to light and reduce the system efficiency. These factors have to be taken into

account when designing a PBR for wastewater treatment.

The algae-bacteria community relevance

As the main “engine” of this treatment, the biological community should be well

understood to understand the treatment itself. However, very little knowledge exists on the

dynamics of the PBRs highly complex communities. Some studies have approach this

matter (Oron et al. 1979, Roeselers et al. 2007, Wantawin et al. 2008), but much more are

needed to understand these interactions between organisms. Knowing the best

performing microalgae in specific conditions, the bacteria that enhance this microalgae

growth and the bacteria enhanced by these microalgae may provide important tools to

optimize the treatment.

Introduction

18

Table 2 - Advantages and disadvantages of the treatment systems for landfill leachate

treatment.

Treatment

technology Advantages Disadvantages

Ph

ys

ic-c

he

mic

al tr

ea

tme

nt

Coagulation-

flocculation High removal rates

Toxic sludge production increase, metals

concentration increase, coagulants cost and

pH sensitivity

Chemical

precipitation

Precipitants low cost, operational

simplicity and valuable struvite

production

Toxic sludge production increase, metals

concentration increase and pH sensitivity

Ammonium

stripping High ammonium removal rates

Only ammonium is removed and costly pH

manipulation

Activated carbon

adsorption High removal rates

Toxic sludge production increase and activated

carbon cost

Chemical

oxidation Oxidizes refractory compounds

Energy demand and toxicity for biological

treatment due to produced oxidized

compounds

Electrochemical High removal rates High cost and quick degradation of the

components

Membrane

filtration Very high removal rates

Membrane fouling and production of high

amounts of highly toxic contaminants

concentrate.

Bio

log

ical tr

eatm

en

t

Leachate transfer Simplicity and low cost Inhibition of the receiving biological treatment

system leading to a series of problems

Aerobic activated

sludge Simplicity

High energy demand, fragility of the biological

community and BOD/COD ratio > 0.5

Aerated lagoons Low cost construction, simplicity and

high removal rates

Big area for construction, odors, algal blooms

and insect infestations.

Trickling filters

Increased resistance of the biological

community, less sludge production and

exempt the use of aeration

BOD/COD ratio > 0.5

Anaerobic

High BOD removal rates, less sludge

production, low P requirements and

methane production

BOD/COD ratio > 0.5, low reaction and

increased fragility due o high ammonium

concentrations and presence of heavy metals

Pure culture

microalgae

High removal rates, no need for

biodegradable organic matter and highly

valuable sludge

Contaminations

High rate algae

pond

High removal rates, simplicity, biological

aeration, no need for biodegradable

organic matter and valuable sludge

Big area for construction

Suspended

biomass vertical

PBRs

High removal rates, biological aeration,

no need for biodegradable organic

matter and valuable sludge

High cost, increased suspended solids in the

effluent and reduced resistance to physic-

chemical variations

Attached biomass

vertical PBRs

High removal rates, biological aeration,

no need for biodegradable organic

matter, valuable sludge, increased

resistance of the biological community

High cost and light penetration

Introduction

19

Objectives

In this thesis a new leachate treatment configuration is proposed. For this

configuration a new PBR of attached biomass was developed. To test the treatment, a

pilot treatment plant was assembled in a landfill facility.

The present study evaluates the efficiency of this new method to remove the high

amounts of ammonium, non-biodegradable COD and other contaminants present in the

landfill leachate. The study also aims to determine the optimum operational factors in

order to make possible the scale up of this system. The phototrophic biofilm community

was also examined in order to increase our knowledge about the dynamic of this biofilm

that forms inside the PBRs.

Chapter I – Pilot leachate treatment plant planning and assembling

20

Chapter I

Pilot leachate treatment plant planning and

assembling


21


The landfill studied in this thesis is located in Ermesinde, Portugal (Figure 3) and

was built by LIPOR. It is a municipal landfill that started to receive urban solid waste on

1970 and was closed on 1995, making this an old landfill. The landfill covers an area of

approximately 19 ha, and the total amount of solid waste deposited is estimated to be

2500000 tons. The leachate from the landfill is conducted to two separate deposits, as

shown in Figure3. The leachate treated in this study was from the deposit marked with the

blue dot and the pilot treatment plant was assembled over the deposit.

Figure 3 - LIPOR I landfill, Ermesinde. Dots show the leachate deposits.

Leachate chemical characterization

The amount of landfill leachate produced is highly variable due to the precipitation

but a 1 m3/h production was estimated based on the annual production registered but this

may increase considerably on rainy seasons. The leachate is more concentrated when

the precipitation is lower and less concentrated when the precipitation is higher. During

the current study, the precipitation was very low. The leachate presents an intense dark

brownish yellow color and very little matter in suspension. Initial analyses made by an

external laboratory showed that the leachate presents high levels of ammonia, between

450 mg/l and 1200 mg/l of NH4-N. The COD and BOD levels ranged from 1100 to 1400

mg/l and 240 to 380 mg/l respectively. The low BOD/COD ratio, approximately 0.24,


22

showed that only a fraction of the organic carbon present in the leachate is biodegradable.

The concentration of phosphorous is also above the legal limits, 1 mg P/l (Decreto-Lei n.o

348/98).

Problems of biological treatment

The high NH4-N concentrations are above the inhibition levels for the AOB and

NOB which are 10-150 mg/l and 0.1-1 mg/l respectively (Kim et al. 2006). The low

biodegradability of the organic carbon may also be a limiting factor for the metabolism of

heterotrophic organisms responsible for nitrification. The low COD/total N ratio, lower than

1.5, is under the necessary ratio for an efficient nitrification, over 4-5 (Pambrun et al.

2008). The NH4-N inhibition values for microalgae development are documented to be 20

mg/l (Borowitzka 1998) but other studies show a good efficiency using microalgae to treat

wastewaters with 405 mg/l (Lin et al. 2007). However this concentration is half the

concentration of the leachate in this study. Plus the intense color of the leachate may also

pose a threat to the effective use of microalgae due to reduced light penetration.

Pilot treatment plant planning

- General plan

The pilot treatment plant consists of four modules: ozone treatment which is

composed by the generator and mixer, trickling filter, laminar clarifier and 9 PBRs with

different diameters (Figure 4). These modules will be better described below.

O3 M

ixe

r

Tri

ck

lin

g

filt

er

Laminar

clarifier

PB

R 2

0

PB

R 3

0

PB

R 4

0

Leachate deposit

O3 Generator

Figure 4 - Pilot treatment plant scheme.


23

- Ozone treatment

The use of ozone (O3) treatment has two main objectives. The first is to oxidize the

ammonium and lower it´s concentration in the leachate and the second is to improve the

biodegradability of the organic matter which will improve the posterior biological treatment.

The O3 treatment system is composed of 2 parts: the generator to produce O3 and the

mixer in which the O3 will be in contact with the leachate (Figure 5). The O3 is dispersed

within the mixer by a venturing pump. This allows a better dispersion of the ozone

enhancing its effect. Before assembling the O3 generator, only air was injected in the

mixer for the treatment. This started to function 2 months before the whole system

assembly. The mixer had been used in a previous urban wastewater treatment plant and

had already some dry biofilm attached. The first generator used produced 300 mg O3/h,

and was assembled in day 14 after the start of the complete treatment system. It

consisted of two generators from Sander, the C100 that produced 100 mg O3/h and the

C200 that produced 200 mg O3/h. On day 47, a new O3 generator was installed and that is

now under operation. This has the capacity to produce 30 g O3/h, inject O3 under pressure

and possesses a cooling system, consisting of a water deposit and a chiller. Foam

produced was collected on the top of the mixer and piped to the leachate deposit.

Foam exit

Leachate exit

Leachate entrance

Venturing pump

Second O3

generator

Air

compressor

Cooling

water

Chiller

First O3

generator

O3 M

ixe

r

Figure 5 - O3 treatment system scheme. Blue arrows – leachate flux; grey arrows –

air/ozone; purple arrows – cooling water; green arrow – foam collection.


24

- Trickling filter with Sycon® plates

A trickling filter was chosen as the first biological treatment for nitrification of the

leachate. The trickling filter is more suitable than suspended biomass technologies due to

the biofilm improved resistance as stated earlier. The objective of this filter is performing

the nitrification process and removing the organic matter. This trickling filter used Sycon®

plates (Figure 6A) as substrate which maximizes the leachate/substrate interface. This

system proved to give good results in previous wastewater treatment plants. The

arrangement of the plates with a rotation of 90o between them improves the contact of the

water with the biofilm (Figure 6B). This arrangement also generates more aeration in the

filter by improved air dragging from the top to the bottom eliminating the need of additional

aeration.

Figure 6 – (A) Sycon® plate; (B) Sycon® plates disposition.

The volume of the filter is 0.342 m3 and the surface area of the Sycon® plates

available for biofilm formation is 59.89 m2. The trickling filter has a small clarifier at its

base and a pump that collects leachate from this clarifier and pumped it to the top of the

filter (Figure 7). This creates a recirculation within the trickling filter that improves the

leachate/biofilm contact time. Due to loss of power during the experiment this pump had to

be replaced. The trickling filter and the Sycon® plates had been used in a previous

wastewater treatment plant and still had dry biofilm attached.


25

E-1

Leachate

entrance

Leachate exit

Sludge exit

Recirculation pump

Figure 7 - Trickling filter scheme. Blue arrows – leachate flux; orange arrows – sludge

collection.

- Laminar clarifier

In order to settle suspended matter, a laminar clarifier was used (Figure 8). The

laminar design improves the settling process and allows the use of a smaller clarifier. It

also allows some biofilm formation that may contribute to the treatment.

Leachate entrance

Leachate entrance Leachate exit

Sludge exit

Figure 8 - Laminar clarifier scheme. Blue arrows – leachate flux; orange arrows – sludge

collection.


26

- New PBR

A vertical column PBR was chosen for the algae-bacteria treatment (Figure 9). An

ascendant flux configuration was used in order to improve the water retention time in the

reactor and also the settling process. The leachate flux is shown in Figure9. To facilitate

the purge of the sludge formed, the base of the reactor was designed leaning towards the

side of the sludge collecting valves.

Due to the advantages of phototrophic biofilms compared to the suspended

phototrophic organisms, transparent polycarbonate alveolar plates were used as substrate

to enhance this biofilm formation (Figure 10A). These were assembled in a vertical

position in the reactors to increase the surface available. The vertical position allows a

better clarification process and prevents clogging effects. The position of the plates inside

the reactor is shown in Figure 10B. The vertical space between the plates has the

objective of allowing the leachate diffusion and increase the light penetration. The reactor

is 2 m high and three different column diameters were used: 20 cm, 30 cm and 40 cm.

Three of each were installed separated by 1 m from each other in the position shown in

Figure 11. The volume and surface area available for biofilm development in the different

PBRs is shown in table 3.

Leachate

entrance

Leachate

exit

Sludge exit

Figure 9 - PBR scheme. Blue arrows – leachate flux; orange arrows – sludge collection.


27

Figure 10 - PBRs substrates. a - polycarbonate alveolar plates; b - position of the

polycarbonate alveolar plates in the PBRs.

Alg 20-1

Alg 20-2

Alg 20-3

Alg 30-1 Alg 40-1

Alg 40-2

Alg 40-3

Alg 30-2

Alg 30-3

Figure 11 – PBRs spatial disposition.

Table 3 - Volume and surface area available for biofilm growth in the PBRs.

PBR

diameter Volume (l) Surface area (m2)

20 cm 55,26 403,34

30 cm 128,74 757,7

40 cm 232,83 1216,36


28

Assemblage

The construction of the leachate treatment plant was phased and at first, in

26/03/2010, only the O3 mixer, trickling filter and clarifier were assembled. In 12/04/2010,

the three 20 cm diameter PBRs were assembled without the transparent polycarbonate

plates and connected to the rest of the system. The other PBRs were assembled and

connected in 05/05/2010 and were also turned on. In 15/05/2010 the transparent

polycarbonate plates were placed inside the PBRs and in 27/05/2010 the flux measuring

system of these was installed. To prevent the PBRs from emptying, anti-return valves

were installed in each PBR. The first ozone generator was installed in 28/05/2010 and the

second was installed in 30/06/2010. The final structure of the leachate treatment plant is

shown in Figure 12 and pictures are shown on Figure 13 and Figure 14.

Circuit description

The leachate is collected in a deposit that already existed under the platform

where the treatment plant was assembled. A pump is placed in the deposit to pump the

leachate. The leachate enters directly in the O3 mixer. In this, in addition to the leachate,

there is also the entrance of treated leachate from the clarifier, creating a recirculation of

leachate within the global system, the big recirculation. After the ozone treatment, the

leachate enters in the lower part of the trickling filter and is pumped to the top. The

leachate is dispersed over the Sycon® plates and trickles by gravity. Following this first

biological treatment, the leachate enters in the clarifier and after clarification; the leachate

is pumped to the big recirculation mentioned above, to the PBR and back into the deposit.

This last portion of the leachate is the treatment plant final effluent. The leachate that goes

to the PBRS enters through the bottom of these and then is collected on the upper part.

From here it is conducted again to the clarifier. The sludge is collected from the bottom

part of the trickling filter, clarifier and PBRs. The leachate flux and sludge collection is

represented in Figure 12.


29

O3 M

ixe

r

Tricklin

g filt

er

Laminar clarifier Ph

oto

bio

rea

cto

r 2

0

Ph

oto

bio

rea

cto

r 3

0

Ph

oto

bio

rea

cto

r 4

0

Leachate deposit

O3 Generator

Air

compressor

ChillerCooling

water

Venturing

pump

Pump

Pump

Pump

Pump

Figure 12 - Leachate treatment plant circuit scheme with second O3 generator. Blue

arrows – leachate flux; orange arrows – sludge collection; grey arrows – air/ozone; purple

arrows – cooling water; green arrow – foam collection.

Figure 13 - Pilot leachate treatment plant with the first O3 generator.


30

Figure 14 - Pilot leachate treatment plant with the second O3 generator.

Operational data

The treatment plant worked in a continuous mode and the leachate entrance flux

was fixed at 50 l/h. The recirculation between the exit of the clarifier and the entrance of

the O3 mixer is 950 l/h, thus diluting the entrance leachate by a factor of 20 with treated

leachate. This has the objective of lowering the concentrations of the leachate entering

the treatment system thus reducing its inhibition effects. The recirculation within the

trickling filter was defined at 2000 l/h. The PBRs flux was defined at 100 l/h.

The O3 production was 300 mg/h when the first generator was used. After the

installation of the second O3 generator, the production was increased gradually from 9 g

O3/h to 24 g O3/h during 6 days. On the day 58 of the experiment, the O3 generator had a

major failure and had to be fixed. Only on day 70 the O3 generation was resumed but

other problems occurred. The big recirculation and PBRs pump had to be replaced and

the venturing pump started to fail. Due to this the O3 diffusion system was changed to

airstones placed inside the O3 mixer. The pilot treatment plant only started to function

properly on day 98.

The sludge collection procedure was conducted twice a week on the day before

the sampling.


31

Table 4 - Hydraulic retention times.

HR

Total treatment 33h 54min

O3 treatment 2,4min

Trickling filter 4,8min

Clarifier 6,6min

PBR 20 cm 36min

PBR 30 cm 1h 18min

PBR 40 cm 2h 18min

System operational problems

The different treatments of the pilot plant were not turned on simultaneously,

having a variation of more than one month between starting times. This complicates the

discussion of the startup of the whole system. In order to minimize this problem, the start

point of the experiment was defined to be the day when the substrate was placed in the

PBRs, day 0. Another drawback for further discussion is the change of the O3 generator

due to the great increase in O3 amounts. The O3 device was changed because the first

one produce very low amounts of O3 for the treatment needed. Problems with the second

O3 generator also occurred. Due to high temperatures, the device had several function

failures, and the O3 production had to be stopped. The O3 generator was not prepared to

function on the outside and the lack of a proper shelter with air conditioner and the

variations in electric current may have contributed to this problem. The device only started

to work properly on day 98. All these problems had effects on the whole system and these

will be taken into consideration when discussing the treatment.

During the experiment there were some operational problems due to pumps failure

and clogging of flux meters which have some effects on the whole treatment. The deposit

pump had to be replaced on day 14 of the experiment, and during one day the treatment

plant was turned off. The trickling filter recirculation pump started to lose power and had

also to be replaced. The recirculation pump had an unknown problem and started to shut

down during the night. The first time this happened the PBR Alg 20 B was half empty and

the Alg 30 B was completely empty due to failure of the anti-return valves. This problem

was solved by removing the biofilm from the valves regularly. The recirculation pump

stopped after this but the PBRs maintained the leachate inside.

Chapter II – Physic-chemical monitoring of the pilot leachate treatment plant startup

32

Chapter II

Physic-chemical monitoring of the pilot leachate

treatment plant startup


33


Material and methods

- Sampling

Parameters measured with probe (pH, temperature and DO) were done in situ by

submerging the probe in the respective location or by extracting a sample to a cup and

measure in locu. The measures were done at least three times a week and the locations

are shown in Figure14.

Samples were taken twice a week for laboratory analysis. These samples were

taken on the places shown in Figure 15. The time between sampling and analysis never

exceeded 6 hours.

O3 M

ixe

r Tricklin

g filt

er

Laminar clarifier Ph

oto

bio

rea

cto

r 2

0

Ph

oto

bio

rea

cto

r 3

0

Ph

oto

bio

rea

cto

r 4

0

Leachate deposit

Pump

Pump

Pump

Alg OUT

Tower

IN

OZ OUT

P-66

Alg20-1 Alg20-3

Alg20-2

Alg30-3

Alg30-2

Alg30-1

Alg40-2

Alg40-3Alg40-1

Figure 15 - Sampling locations.

- In situ measurements

The pH, temperature and DO were measured in situ using portable meters,

HI991300 and HI9146 from HANNA and a PCD650 from Eutech.


34

- Biological oxygen demand measurement

The 5 day BOD was analyzed following the standard methods (Greenberg et al.

1992). The DO was measures using the HI9146 meter from HANNA.

- Chemical oxygen demand, total nitrogen, ammonium, nitrites, nitrates and total

phosphate measurements

These parameters were measured spectrophotometrically with commercial kits

from HACH-Lange (Table 5). All the procedures were done according to the manufacturer

instructions and using a reactor HT200S and a DR2800 spectrophotometer from HACH-

Lange. For some tests, dilution of the sample was needed so that results fell within the

range of the test. A commercial kit, Addista from HACH-Lange, was used every two

weeks as a control to verify any interference in the analytical procedures.

Table 5 - HACH-Lange kits used for chemical analysis.

Parameter Kit used Range

COD LCK514 100 – 2.000 mg/l

Total nitrogen (Total N) LCK338 20 –100 mg/l TNb

Ammonium LCK302 47– 130 mg/l NH4-N

Nitrite LCK342 0,6 – 6 mg/l NO2-N

Nitrate LCK340 5 – 35 mg/l NO3-N

Total phosphate LCK350 2 – 20 mg/l PO4-P

Iron LCK321 0,2 – 6 mg/l

Results and discussion

- Initial remarks

The first oxygen probe presented problems of calibration, and the DO results had

wide variations. This interfered with the BOD quantifications and the results were not

considered. The only BOD values mentioned were measured with the first oxygen probe

but only when this was used exclusively in the laboratory which eliminated the calibration

problems.

The analysis of the results is separated between the use of the first O3 generator

and the second one. Due to major failure of the second generator and following problems,

only the results recorded after the restart of the O3 generator were used. This was done to

eliminate the variables introduced by the O3 treatment break and following problems and


35

allow the discussion of more consistent values. The results with the second O3 generator

and their comparison with those of the first O3 generator are referred on the final topic of

this chapter.

- Untreated leachate

The untreated leachate composition changed during the study (Figure 16). This

happens due to phenomena inherent to landfills dynamics.

Figure 16 – Untreated leachate (A) pH, temperature; (B) COD, total N, NH4-N; (C) PO4-P

and Fe variation.

0

10

20

30

6,5

7

7,5

8

8,5

1 5 9 13 17 21 25 29 33 37 41 45 49

oC

pH

Days

pH Temperature

600

800

1000

1200

1400

1600

1800

3 6 10 13 17 20 24 27 31 34 38 41 45 48 52

mg/

L

Days

Total N NH4-N COD

0

5

10

15

20

0

20

40

60

80

3 6 10 13 17 20 24 27 31 34 38 41 45 48 52

mg

Fe/L

mg

PO

4-P

/L

Days

PO4-P Fe

A

B

C


36

The leachate temperature was almost always between 20 and 25 oC and pH was

always around 7 and 7.2 which shows that these parameters were quite stable. The COD

levels ranged between approximately 1200 and 1350 mg/L and the BOD/COD ratio was

rather low, 0.1 or lower, being consistent with the values expected due to the landfill age

(Huo et al. 2008). The NH4-N levels increased during the experiment, from approximately

900 to 1200 mg/L, which is explained by the lack of precipitation, which leads to an

increase in leachate concentration. The total N values also followed this tendency but

presented some punctual increases. Total N concentrations measured on days 10 and 17,

were lower than the NH4-N concentrations. This indicates a possible interference by the

leachate composition in the total N or NH4-N analytical methods. The NO2-N and NO3-N

levels were always low (data not shown) being the maximum concentrations recorded 2.1

and 8.27 mg/L respectively. The Fe concentrations increased during the first 20 days and

stabilized after this. The PO4-P concentrations decreased at the beginning of the study to

around 10mg PO4-P/L, rose punctually on day 31, decreased again to the same values

and increased again on day 52. The suspended solids were always below 0.1 mg/L which

is a very low value and is within the legal limits in Portugal.

- Ozone treatment

The O3 treatment influent is 5% of the untreated leachate (IN) and 95% of the

treated leachate (OUT). The treatment with the first ozone generator had a reduced effect

in the leachate quality and did not accomplish the expected objectives, namely the

reduction of COD and NH4-N and increase of the COD/BOD ratio. Only the NO2 and NO3

produced by the biological treatment were removed (Figure 17). At the end of this study

the treatment was still improving, showing that the system didn’t reach the maximum

performance. Further analysis should be performed to evaluate the maximum

performance.

Figure 17 – Leachate NO2-N and NO3-N concentrations before and after the O3 treatment.

0

50

100

150

200

20 24 27 31 34 38 41

mg/

L

Days

NO2-N - Ozone entrance NO2-N - Ozone exit

NO3-N - Ozone entrance NO3-N - Ozone exit


37

Even though the treatment had a low efficiency, it contributed to the aeration of the

leachate increasing the DO (Figure 18) which enhances the posterior biological treatment

due to promotion of aerobic organisms growth.

Figure 18 – Leachate DO before and after the O3 treatment.

- Trickling filter treatment

With this treatment it was expected to reduce COD, BOD and NH4-N levels.

However, the achieved COD and NH4-N reductions were quite low (Figure 19). The

treatment had already been working for more than 1 month before monitoring, so it was

expected to provide already good efficiencies which was not the case.

COD removal was not observed but on the contrary, there was a slight increase in

COD after this treatment maybe due to the decomposition of the biomass produced in the

reactor. On days 34 and 41 the removal rates decreased to negative values, showing a

slight production of COD. On day 34, BOD concentrations increased slightly, 6.2 mg/L,

while on day 41 BOD concentrations were reduced, 24.6 mg/L. The BOD result from day

34 is inconsistent with the increase of total N, NH4-N and PO4-P removal verified on this

day. This removal indicates that heterotrophic bacteria are active and during their activity

BOD is consumed (Xu et al. 2010). Therefore, the BOD concentration should decrease

instead of increase as was observed. The possible explanation for this inconsistency may

relay on an analytical error of the BOD quantification due to problems above referred.

The NH4-N removal was low, increasing to a maximum of 13% on day 34. The lack

of removal is explained by the BOD/NH4-N ratio. Whatever the ratio at the entrance, the

leachate BOD/NH4-N ratio is always around 0.05 at the trickling filter exit, demonstrating

that NH4-N removal is limited by the BOD concentrations. Below this ratio, NH4-N is not

removed. On other studies glucose or methanol were added to the treatment in order to

improve this ratio and increase nitrification efficiency (Visvanathan et al. 2007,

-1,00

1,00

3,00

5,00

7,00

9,00

11,00

13,00

15,00

17 20 23 26 29 32 35 38 41 44

mg/

L

Days

DO - Ozone entrance DO - Ozone exit


38

Tsilogeorgis et al. 2008). However in the current study this was not done because it was

expected that the O3 treatment had improved the biodegradable organic matter content.

The total N removal rate was not constant, and wide variations occurred during the

experiment. Both decreases and increases in total N concentrations were observed and

cannot be explained by the NH4-N, NO2-N and NO3-N variations (Figure 19 and 20). This

indicates that the variations may be related to other nitrogen compounds. The maximum

total N increase was observed on day 27 which is after an increase of total N in the

untreated leachate verified on day 24. This indicates that possibly the total N was

accumulated in the biomass of the trickling filter and its detachment and permanence in

suspension explains the increase in total N.

Figure 19 – COD, total N, NH4-N and PO4-P removal rates with the trickling filter

treatment.

Production of NO2-N and NO3-N was observed mainly on and after day 34 (Figure

20). For both cases the production rate increased from 22.2% and 12% on day 27 to

97.6% and 115% on day 34 respectively. This production increase matches the

ammonium removal increase verified on day 34, showing that nitrification process is

taking place. However, the high amounts of NO2-N indicate that the nitrification is partly

incomplete possibly due to inhibition of the nitrite-oxidizing bacteria that are inhibited by

lower ammonium concentrations than ammonia-oxidizing bacteria are. It is possible that

the ammonia-oxidizing bacteria are degrading NH4-N to NO2-N, while the nitrite oxidizing

bacteria are inhibited and do not degrade the NO2-N to NO3-N, resulting in an

accumulation of NO2-N as observed in this study (Bai et al. 2009).

-15,00%

-10,00%

-5,00%

0,00%

5,00%

10,00%

15,00%

20,00%

6 20 27 34 41

rem

ova

l rat

e

Days

COD Total N NH4-N PO4-P


39

Figure 20 – Leachate NO2-N and NO3-N concentrations before and after the trickling filter

treatment.

- Clarifier treatment

The main objective of the clarifier was to remove TSS which were already below

the legal limits in Portugal. However, the clarifier also removed some compounds and

increased concentrations of others (Figure 21). This happened because of the

phototrophic biofilm formation on the walls and laminas. Only the COD levels were not

affected by the clarifier.

NH4-N was constantly reduced by approximately 10%, and a slight removal increase in

time was observed. The total N was either produced or removed throughout the

experiment. The increase observed on day 41 shows a high increase in organic forms of

nitrogen within the clarifier due to the low variation of NH4-N, NO2-N and NO3-N levels.

The NO2-N and NO3-N concentrations (Figure 22) started to increase slightly at the

beginning of the experiment but on day 34 they started to be reduced.

The PO4-P was also removed and produced in the clarifier during the experiment

(Figure 21). On day 34 there was a high peak of PO4-P production, 55%, that may have

contributed to the removal increase peak verified in the PBRs (Figure 23). The cause of

this production peak is related to the high PO4-P peak observed in the untreated leachate

on day 31 (Figure 16C). This high amount of PO4-P was incorporated in the biomass

during the treatment and this last accumulated in the clarifier as supposed to. The high

production peak verified in the clarifier could be due to the decomposition of this P-rich

accumulated biomass that may have been inefficiently removed during the purge process

that took place one day before de analysis.

0

50

100

150

200

250

3 6 10 13 17 20 24 27 31 34 38 41 45

mg/

L

Days

NO2-N - Tower entrance NO2-N - Tower exit

NO3-N - Tower entrance NO3-N - Tower exit


40

Figure 21 – COD, total N, NH4-N and PO4-P removal rates by the clarifier.

Figure 22 – NO2-N and NO3-N removal rates by the clarifier.

- PBRs treatment

The PBRs objective was to remove mainly the NH4-N and COD from the leachate

but also other contaminants such as NO2-N, NO3-N, PO4-P and iron and the results of the

overall PBRs system are shown in Figures 23. Three PBRs of each diameter were built in

order to function as replicates, but environmental factors such as sunlight incidence or

wind may affect the PBRs differently depending on their spatial positioning. Variations

between the three replicates were taken into account when analyzing the physic-chemical

results that are shown in Figure 23. For all parameters evaluated small variations were

observed between the three different diameters.

The COD removal was low, normally below 2%, achieving maximum reduction

values on day 34 of 3%, 4.2% and 4.5% for the Alg 20, Alg 30 and Alg 40 PBRs

respectively.

The total N removal was not constant (Figure 23) as seen for the other treatments.

Low values, above 10% of production, were recorded on the days 10, 17, 31 and 34; and

high values, above 20% of removal, on the days 24 and 41. As in the trickling filter

treatment, these variations cannot be explained by the inorganic nitrogen levels. On day

24, which corresponds to the highest removal values, the untreated leachate presented an

-60%-50%-40%-30%-20%-10%

0%10%20%

20 27 34 41

rem

ova

l rat

e

Days

COD

Total N

NH4-N

PO4-P

-20%

-10%

0%

10%

20%

30%

20 27 34 41

rem

ova

l rat

e

Days

NO2-N

NO3-N


41

increase in total N amounts that was also not due to the NH4-N, NO2-N and NO3-N

amounts. This denotes a possible relation between the two events and indicates that this

must be related to the increase of organic nitrogen compounds not measured in this

study. The NH4-N removal was quite constant and low, achieving a maximum of 4.4% on

day 24 (Figure 23). This removal developed equally for the three PBRs but on day 41, a

differentiation appeared (Figure 24) and, even though significant variations have been

obtained between the replicates, the Alg 30 was the most efficient followed by the Alg 20

and the Alg 40. NO2-N and NO3-N suffered small variations between the leachate entering

in the PBRs and the one leaving from these. During growth, algae incorporate NH4-N and

do not degrade it like in nitrification process. Consequently NO2-N and NO3-N levels

should not increase. Microalgae should also incorporate the NO3-N and reduce its

concentrations (Fierro et al. 2008). In these PBRs, aerobic bacteria were also growing

inside and nitrification should have been noticed. A possible explanation for the reduced

variation in NO2-N and NO3-N concentrations is that equilibrium was established between

microalgae and bacteria. Ammonia oxidizing bacteria and nitrite oxidizing bacteria

produced these compounds and microalgae present consumed them, contributing to the

NH4-N removal without changing the NO2-N and NO3-N concentrations considerably.

Figure 23 - COD, total N, NH4-N and PO4-P removal rates by the whole PBR system.

Efficient PO4-P removal (Figure 23) was only observed on day 31, which followed

a peak of PO4-P in the untreated leachate (Figure 16C) and coincided with the production

increase in the clarifier on day 34 (Figure 21). The possible reason for this event was

already mentioned before in the clarifier results. This event could also be the cause of the

high production of total N on day 31 due to the increase of biomass growth and

subsequent increase of organic nitrogen. On this day, Alg 40 performed better, with a

removal rate of 38.3%, followed by the Alg 20 and the Alg 30, with 30.9% and 26.6%

respectively (Figure 24). On the other days, PO4-P removal rate was low, always under

5%. This suggests that a higher PO4-P concentration may enhance the treatment process

-30%

-20%

-10%

0%

10%

20%

30%

40%

50%

3 6 10 13 17 20 24 27 31 34 38 41 45

rem

ova

l rat

e

Days

COD

Total N

NH4-N

PO4-P


42

possibly due to an increase of biomass growth, but only with further studies this could be

confirmed. Other studies have shown that biomass and phosphorous levels are related

and an increase in phosphorous leads to an increase in biomass (Guzzon et al. 2008).

Figure 24 – (A) COD, (B) total N, (C) NH4-N and (D) PO4-P average removal rates from

the three replicates of the different PBRs.

-2,0%-1,0%0,0%1,0%2,0%3,0%4,0%5,0%6,0%

6 20 27 34 41

CO

D r

em

ova

l rat

e

Days

Alg 20

Alg 30

Alg 40

-30,0%

-20,0%

-10,0%

0,0%

10,0%

20,0%

30,0%

40,0%

6 20 27 34 41

Tota

l N r

em

ova

l rat

e

Days

Alg 20

Alg 30

Alg 40

-15,0%-10,0%

-5,0%0,0%5,0%

10,0%15,0%20,0%25,0%

6 20 27 34 41

NH

4-N

re

mo

val r

ate

Days

Alg 20

Alg 30

Alg 40

-10,0%

0,0%

10,0%

20,0%

30,0%

40,0%

50,0%

6 20 27 34 41

PO

4-P

re

mo

val r

ate

Days

Alg 20

Alg 30

Alg 40

A

B

C

D


43

The PBR treatment generated a slight increase in pH (Figure 25), which may be

due to microalgae growth as observed by other authors (Schumacher et al. 2003).

Temperature has sometimes increased probably due to the heat up the PBRs by sunlight

incidence.

Figure 25 – Leachate pH and temperature variation before and after the PBR treatment.

- Pilot treatment plant results

The overall results of the treatment plant with the first O3 producer showed that this

configuration was not capable of treating the leachate with a significant efficiency (Figure

26), as the values of the legal parameters in the treated leachate are far above the legal

limits. To achieve these values, removal rates of approximately 90%, 99% and 87% for

COD, total N and PO4-P respectively should be reached.

Figure 26 - COD, total N, NH4-N and PO4-P removal rates by the pilot treatment plant.

20222426283032343638

7,3

7,4

7,5

7,6

7,7

7,8

7,9

8

10 13 16 19 22 25 28 31 34 37 40 43 46

oC

pH

Days

ph - PBRs entrance pH - PBRs exit

Temperature - PBRs entrance Temperature - PBRs exit

-10%

0%

10%

20%

30%

40%

50%

3 6 10 13 17 20 24 27 31 34 38 41 45

rem

ova

l rat

e

Days

COD

Total N

NH4-N

PO4-P


44

Figure 27 – Leachate (A) COD, (B) total N, (C) NH4-N, (D) NO2-N, (E) NO3-N, (F) PO4-P,

(G) Fe, (H) pH, (I) temperature and (J) DO evolution along the treatment plant and time.

IN dil 20x concentration is the calculated value of the leachate entering in the O3 mixer. It

correspond to a 20x dilution of the IN leachate with the OUT leachate.

A B

C D

E F

G H

J I


45

The COD concentrations were barely reduced, approximately 10 %, and this was

majorly achieved by the O3 treatment (Figure 27A). The biological treatments did not show

a constant performance, exhibiting COD reduction or production on different days. On

days 34 and 41, the COD was increased in the trickling filter treatment possibly due to

degrading sludge that was not removed efficiently during the purge protocol. The BOD

levels were reduced by more than 50%, achieving a maximum removal rate of 93.9% on

day 20 and reducing after this to a minimum of 54.7% removal on day 41. Even though

low amounts of BOD were available, these were not fully consumed by the aerobic

bacteria as expected. This may be explained by the inhibition of the aerobic bacteria by

other factors, such as high NH4-N levels, which leads to a lower usage of the organic

matter by these organisms.

Total N removal was very variable and a maximum of 25.1% were observed on

day 45 (Figure 26). This removal seemed to be accomplished mainly by the trickling filter

with the exception of the day 27, when the total N removal rate increased but the trickling

filter contributed negatively to it due to reasons discussed before (Figure 27B). The other

treatments showed wide variable results making it difficult to draw conclusions about their

role in the total N concentrations. Removal rates of NH4-N were also not constant. These

were between 10% and 20% most of the time, and values of 40.1% and 22.8% occurred

on days 17 and 41 (Figure 26). This removal was mostly done by the trickling filter (Figure

27C) and only a small part by the clarifier and PBRs. The PBRs should have performed

better NH4-N removal efficiencies than the trickling filter because BOD was low for the

efficient functioning of the last. Once more, inhibition effects may have occurred due to the

leachate high contamination. The trickling filter and PBRs present different microbial

communities with different susceptibilities to contamination. As an example, the growth of

microalgae present in the PBRs is inhibited by lower concentrations of NH4-N than

ammonia oxidizing bacteria, more commonly present in the trickling filter. The NH4-N

degradation within the trickling filter led to an increase in NO2-N and NO3-N as already

discussed before (Figure 20). The trickling filter was the main responsible for producing

these compounds which were removed mainly by the O3 treatment but also in smaller

rates by the PBRs (Figure 27D and 27E) and clarifier.

The PO4-P removal rates were the highest compared to the other compounds

removal (Figure 26). The amounts of PO4-P in the untreated leachate were low and

suffered wide variations as already discussed, which indicates that PO4-P was a limiting

factor. It was noticed that PO4-P removal trend is similar to the NH4-N trend, showing a

relation between their degradation. It is known that increasing the PO4-P may lead to

better leachate treatment due to a better C:N:P ratio. In other studies, additional PO4-P

was added to the system to improve treatment (Tsilogeorgis et al. 2008). All the biological


46

treatments contributed to PO4-P removal (Figure 27F), and this was verified more

significantly on day 34, when a peak in PO4-P concentrations occurred.

Fe concentrations (Figure 27G) were also reduced by the treatment plant, mainly

by the PBRs where microalgae use Fe as a nutrient for their growth.

The pH values (Figure 27H) increased throughout the treatment plant after a

reduction in the O3 treatment, being this increase most significant in the PBRs as

expected. This and the DO increase (Figure 27J) in the PBRs indicate that phototrophic

activity is being performed inside the PBRs. DO was increased initially by the O3 treatment

and was reduced in the following trickling filter and clarifier due to consumption by the

aerobic bacteria. In the PBRs a DO increase was observed after day 24, showing that

microalgae were growing and phototrophically active. On day 45, DO decreased in the

PBRs which indicates that heterotrophic bacteria were growing actively and consuming

the O2 produced by microalgae. This will be further discussed on the microbial community

study.

The PBRs were also the major contributor to the temperature increase (Figure 27I)

which is due to the heat up of the reactors as discussed before. An initial temperature

decrease was achieved by the O3 treatment due to insertion of gas at outside temperature

that promoted a cooling effect on the leachate.

The overall treatment showed low efficiency when using the first O3 generator. The

installation of the second O3 generator is expected to produce better results by efficiently

removing NH4-N from the leachate thus increasing its biodegradability. The lower levels of

NH4-N should diminish the inhibition effect on the microbial communities, mainly the

microalgae, and the improved biodegradability may enhance nitrification in the trickling

filter.

- Treatment comparison between first and second O3 generators

Average values from two samples of the first O3 generator, days 34 and 41,

treatment and two samples of the second O3 generator, days 101 and 103, were used to

compare the treatments. When the whole treatment was analyzed, differences were

noticeable between the two conditions (Figure 28).

It is clear that COD removal in the overall system is significantly more efficient

with the second O3 generator. Due to the higher amounts of O3, the organic carbon

compounds may be oxidized to more biodegradable and less toxic forms, thus improving

the efficiency of the biological leachate treatment. The total N and NH4-N removals

differences were not significant.


47

Figure 28 – Comparison of COD, total N and NH4-N removal percentages by the whole

treatment between the two O3 generators.

Figure 29 - Comparison of (A) NO2-N and (B) NO3-N production by the whole treatment

between the two O3 generators.

Another difference detected between the two settings was in the NO2-N and NO3-N

results during the whole treatment (Figure 29). The amounts of NO2-N (Figure 29A) were

significantly lower on the second O3 generator setting than on the first and the amounts of

NO3-N (Figure 29B) had an opposite result. This shows that with the second O3 generator,

nitrification process is increased, allowing a reduction in NO2-N and consequent increase

in NO3-N. Consistently, the NH4-N concentrations also reduced which favors the

nitrification process.

0%

5%

10%

15%

20%

25%

30%

COD Total N NH4-N

Re

mo

val r

ate

First O3 generator

Second O3 generator

0

50

100

150

200

250

First O3 generator

Second O3 generator

mg

NO

2-N

/L

IN

OUT

0

10

20

30

40

50

60

70

First O3 generator

Second O3 generator

mg

NO

3-N

/L

IN

OUT

A B

Chapter III – Photobioreactors biofilm characterization

48

Chapter III

Photobioreactors biofilm characterization


49


Material and methods

- Sampling

To study the biofilm development in the PBRs, small polycarbonate plates of 1 cm2

were fixed to a string and were placed inside the PBR Alg 40-2 (Figure 30) in the center

(C) and on the periphery (P) of the PBR at approximately 0.5 and 1.5 m from the top

(Table 6). The plates were collected on the following days of the experiment: 0, 3, 6, 13,

20, 27, 34, 41 and 48, and were frozen at -80oC inside sterile sampling tubes.

Figure 30 - Polycarbonate plates of 1 cm2 placed inside the PBR (a) before filling the PBR,

(b) after filling the PBR, and (c) 4 days after functioning.

Table 6 – Positioning of the sampling plates inside the PBR. C – center; P – periphery.

Sample Position

0.5C PBR center at 0.5 m from the top

1.5C PBR center at 1.5 m from the top

0.5P PBR periphery at 0.5 m from the top

1.5P PBR periphery at 1.5 m from the top


50

- Sampling plates optical analysis

The plates were observed with a magnifier LEICA GZ4 and optical microscope

LEICA CME. A portion of the biofilm on the plates was removed and observed under an

optical microscope. Images of the plates were acquired with a photographic camera

(colored images) and also with the GenoSmart system from VWR (black and white

images).

The images were analyzed with ImageJ software to quantify the coverage

percentage of the phototrophic biofilm and the mean gray value and to evaluate the

biofilm thickness profiles. The degree of phototrophic biofilm colonization of the plates was

estimated by splitting the image by RGB color layers. The quantification was done using

only the green layer. For this layer the threshold was defined to mark the phototrophic

biofilm and then the marked area was measured as area percentage. The mean gray

value and its standard deviation were measured from the images taken with the

GenoSmart system. Biofilm thickness profiles were assessed using the 3D surface plots

function with the colored images of the sampling plates.

- Microalgae identification and quantification

Microalgae were identified morphologically by microscopic observation. For

microalgae quantification, the biofilm was removed with the help of a sterile toothpick,

placed in 3 ml of distilled water and submitted to vortex. Cell number was assessed by

microalgae counting in a Sedgewick Rafter chamber and expressed as microalgae

cells/cm2. Before the counting procedure, the sample was left to settle 3 minutes in the

chamber. This was done in triplicate for all samples.

- Heterotrophic bacteria

The heterotrophic bacteria were quantified by the spread plate method in NWRI

agar (HPCA) medium (Greenberg et al. 1992). The optimum dilution used in the

procedure was determined previously by testing various dilutions of two samples and

choosing the one that produced results between 30 and 300 colony forming units (CFUs).

The biofilm was extracted from one side of the plate with a sterile razor blade,

placed in 1 ml sterile distilled water and vortexed for 1 minute. The samples were diluted

with the dilution factor determined previously. Then 0.3 ml of the diluted sample were

placed on the NWRI medium in the petri dish and the sample was distributed over the

media surface with a glass rod. Cultures were incubated for 5 days at 26 oC in an inverted

position and then CFUs were counted and the value expressed in CFUs/cm2. This

procedure was done for three replicates of each sample.


51

- Dominant microalgae isolation

Cultures were done as previously described for the heterotrophic bacteria analysis

but the inoculated plates were incubated at room temperature and exposed to light. Green

colonies sub-cultured on a new plate with NWRI agar (HPCA) medium using the streak

plate method. Single colonies were picked for microscopic observation and as an

inoculum for new culture on a plate with NWRI agar (HPCA) medium.

Results and discussion

- Initial notes

The sampling plates did not stay all in the same position which caused differences

in biofilm formation within the same sample group. When the samples were collected, the

dragging from inside the PBRs may have caused some biofilm detachment. The 0.5P

sampling plates were lost after the sampling day 27. These problems will be taken into

consideration in the following results discussion.

- In situ observations

The development of phototrophic biofilm on the PBRs wall was assessed by direct

observation from the outside of the bioreactor and from the inside after emptying slightly

the PBR. The appearance from the outside changed with time as seen in Figure 31.

After only one day, the PBRs wall already exhibited an almost transparent thin

layer of biofilm that was only visible without leachate in the PBRs. This initial biofilm

generates better conditions for the attachment of microalgae as has been shown in a

previous study where heterotrophic bacteria were responsible for a faster development of

the phototrophic biofilm (Roeselers et al. 2007). During the following days, small colonies

of microalgae with a light green color started to appear. These expanded and on the 4th

week the PBRs walls were completely covered by a light green layer. After this point, the

biofilm started to darken and had a more compact appearance. On the 6th week the color

started to change to a dark greenish brown while, from the inside, the biofilm presented a

dark green color. This suggests a differentiation in metabolism between biofilms in the

wall and leachate interfaces, showing a clear stratification. Such effect may be due to lack

of contact of the biofilm-wall layer with the leachate and subsequent lack of nutrients or to

the high amounts of DO produced by the biofilm-leachate layer that act as an oxidant. The

greenish brown color started to change to a lighter color during the transition to the 7th

week, suggesting an increase of heterotrophic bacteria. By this time, the biofilm started to

present small regions in the biofilm-wall layer that consisted of gas bubbles (Figure 32)


52

which are usually associated with anaerobic bacteria activity (Wang 2009). This provoked

the detachment of biofilm by the 8th week, opening space for new phototrophic biofilm

development which was already grown one week after. It was also noticed that when

many detachment regions start to appear, bigger portion of biofilm got detached as shown

in Figure 32. This phenomena increases light penetration to the inner alveolar plates,

increasing the growth of the phototrophic biofilm on these.

Figure 31 - Schematic representation of the biofilm formed on the walls of the PBRs,

observed from the outside.

The biofilm formation in the PBRs wall presented the expected dynamics, taking

approximately 9 weeks to stabilize. The stability consists of a cycle of phototrophic biofilm

growth, aging and detachment that lasts approximately 5 weeks.

The phototrophic biofilm formation on the PBRs wall may be counterproductive

since it decreases the light that reaches the alveolar plates inside the PBRs. Future

designs should have this aspect into consideration.


53

Figure 32 – Biofilm detachment in the PBRs: (A) small detachment spots and (B) big

detachment portions.

- Sampling plates optical analysis

The biofilm formation on the sampling plates is shown in Figure 33. The first

phototrophic biofilm started to appear on day 6 on the peripheral plates. This biofilm had a

light green color which got darker as the biofilm started to get thicker. Microscopic

observations show that the dominant organism in this biofilm is a microalga, Chlorella sp.,

which was the main responsible for the biofilm color. It was already shown that Chlorella is

among the most resistant microalgae in environments with high organic contamination

(Gonzalez et al. 1997; de Godos et al. 2010). The phototrophic biofilm expansion was

slightly quicker on the 0.5P than on the 1.5P samples, reaching the 100% plate coverage

on day 20 and on day 34 respectively (Figure 34). This difference may be explained due

to a better access to light by the 0.5P plates than the 1.5P ones. It is also noticeable that

on day 48, the 1.5P sample had a darker biofilm with a slightly brownish color, suggesting

development of other microorganisms or formation of EPS with this color.

On the inner plates, a pigmented biofilm development, consisting of small colonies,

was observed after day 6 but only after day 20 this biofilm expanded considerably. On the

0.5C plates, even though the biofilm presented a great number of microalgae cells, they

present a very light green color, possibly denoting a weak phototrophic activity due to the

lower light levels reaching the center of the PBR. These microalgae formed clusters that

generated a light brown biofilm over them, mainly composed of bacteria and EPS, as

observed microscopically. On the 1.5C plates similar results were observed but the biofilm

pigmentation was much weaker and the clusters formed presented a slight darker

brownish color. The microscopic observations showed less microalgae and many with

A B


54

very light green color which was expected due to the even more reduced access to light.

The biofilm pigmentation in the inner samples showed a curious dynamics, increasing and

reducing between weeks with no known cause (Figure 34). This biofilm is quite resistant

to shearing so the loss of biofilm during the samples collection does not explain these

results nor does the different position of the sampling plates as all plates within a sample

group were equally colonized. Further studies with more samples could shed light on this

process.

Figure 33 – Color images of the sampling plates.

Figure 34 – Evolution of the pigmented biofilm coverage on the sampling plates.

The biofilm thickness was evaluated using the color of the biofilm. Darker biofilm

regions were thicker than the other due to higher amounts of pigmented cells and EPS

leading to a bigger absorption of light.

0%

20%

40%

60%

80%

100%

0 3 6 14 20 27 34 41 48

Pig

me

nte

d b

iofi

lm c

ove

rage

Days

0,5C 1,5C 0,5P 1,5P


55

Biofilm thickness profile of the sampling plates, Figure 35, shows that higher

thickness levels were only developed on the peripheral samples after day 6 on 0.5P

samples and after day 14 on 1.5P samples. On the 1.5P, an increasing more

homogeneous thick biofilm was observed on day 27 and following samples,

corresponding to weeks 4 to 7. These results are in agreement with the in situ

observations of the PBR wall (Figure 31). Unfortunately and due to lack of samples, it was

not possible to confirm the results obtained on weeks 8 and 9 schematized in Figure 31.

The thick regions formed first, as observed on the 1.5P on day 14, suggest that the

increase of biofilm thickness promotes the growth of upper layers of microalgae possibly

due to the substratum properties and production of growth promoting compounds by the

biofilm present. This may be considered as an exponential growing phase of the biofilm

that has been already been shown (Roeselers et al. 2007). Thick regions present on the

1.5P samples on day 14 were quite fragile and very susceptible to shear forces, what was

shown by the easy detachment during the analysis procedures. This did not happen on

samples after day 14, showing that these samples had a more shear resistant biofilm on

thick biofilm regions.

Figure 35 - Biofilm thickness profiles of the sampling plates.

The biofilm thickness on the inner plates also shows the curious dynamic

mentioned above, being this clear in 0.5C samples where medium thickness levels on day

20, 34 and 48, were alternated with low thickness levels on days 27 and 41. Microscopic


56

observation confirmed that this thicker and pigmented biofilm is consisted of small clusters

of microalgae and high amounts of EPS with a brownish color.

Non-pigmented regions of the plates presented a thin biofilm that was able to

reflect light as shown in Figures 36 and 37. This is more noticeable on samples 1.5C, in

which the pigmented biofilm did not develop considerably and the light reflection had the

highest mean gray values (Figure 37), showing that the light reflecting phenomena is

related to the non-pigmented biofilm. Pigmented biofilm regions appear darker due to light

absorption. The samples 1.5P presented an exponential maximum development of the

non-pigmented biofilm until day 6, continuing to increase until day 20 after which the

reflection levels decreased. This decrease is consistent with the increase of the

pigmented biofilm (Figures 33 and 34). All the other samples presented a 3 days lag

phase in the development of the biofilm. The inner samples reached higher mean gray

values than the peripheral samples. On day 48 the mean gray values decreased on the

0.5C sample possibly reflecting the increase of pigmented biofilm as confirmed by the

increase the thickness profiles (Figure 35) and microalgae cell counts (Figure 38).

However, the light reflected are not consistent with the decreases of biofilm thickness and

pigmentation on days 27 and 41.

Figure 36 – Black and white images of the sampling plates.

The increase in light reflection during the experiment may also indicate that the

non-pigmented biofilm was increasing in density because there was no increase in biofilm

thickness as observed microscopically. This biofilm may function as a barrier to light but

the reflecting properties may also be considered an advantage. The light reflected may be

used by other biofilm layers increasing the phototrophic activity.


57

Figure 37 – Evolution of the mean gray values of the sampling plates.

- Microalgae and heterotrophic bacteria dynamics

Microalgae development in the biofilm presented an initial exponential phase

attained first on the peripheral samples and subsequent stationary phase (Figure 38). The

highest number of cells was obtained on the 1.5P, 4.31x106 on day 41. The 0.5P samples

reached 1.5x106 microalgae cells/cm2 on day 27 and after this day no more samples were

available.

Microalgae from these deeper zones consume CO2 and nutrients decreasing their

amounts for the microalgae growing in the upper zones, thus affecting their growth rate.

An increase in flux could increase the amounts of these compounds entering in the PBRs

and contribute to a faster increase of microalgae cells number and subsequently of

biomass, which is beneficial for the leachate treatment. External CO2 supply could also

increase the microalgae growth, thus improving the treatment.

The viable heterotrophic bacteria concentration (Figure 39) shows a high increase

on the peripheral samples. On day 14 samples 0.5P and 1.5P reached a maximum of

1.45x105 and 8.08x104 CFU/cm2 respectively. By day 27, the CFU number decreased to

1x102 and 4.2x103 on 0.5P and 1.5P samples respectively. This reduction matches the

high degree of phototrophic biofilm colonization (Figure 33) and high microalgae number

(Figure 38) that occurred, showing a possible relation between these two events. The

microalgae increase may lead to production of toxic compounds to the heterotrophic

bacteria community present until this day. Heterotrophic bacteria counts recovered to

3.02x104 on day 34 (data only available for 1.5P samples) and kept approximately stable

0

50

100

150

200

250

0 3 6 14 20 27 34 41 48

Me

an g

ray

valu

e

Days

0,5C

1,5C

0,5P

1,5P


58

until the end of the study. This possibly indicates a change in the bacterial community

composition, being this community able to co-exist with the microalgae.

Figure 38 – Evolution of the microalgae cells/cm2 during the experiment.

Figure 39 – Evolution of the heterotrophic bacteria during the experiment expressed in

CFU/cm2.

In the inner samples both microalgae and viable heterotrophic bacteria developed

slower than in the peripheral samples and also reached lower concentrations. This lower

biofilme developing rate on the inner samples is in agreement with the results obtained

with other methods (Figures 33, 35 and 36). The 0.5C and 1.5C samples, had microalgae

1,0E+00

1,0E+01

1,0E+02

1,0E+03

1,0E+04

1,0E+05

1,0E+06

1,0E+07

3 6 14 20 27 34 41 48

log

Mic

roal

gae

ce

lls/c

m2

Days

0,5C

1,5C

0,5P

1,5P

1,E+00

1,E+01

1,E+02

1,E+03

1,E+04

1,E+05

1,E+06

3 6 14 20 27 34 41 48

CFU

/cm

2

Days

0,5C

1,5C

0,5P

1,5P


59

counts almost always under 5x105, being the only exception the sample 0.5C on day 48.

The difference of microalgae increase rate between 0.5C and 1.5C samples is explained

due to a better access to light by the 0.5C because these samples have access to light

that enters through the top of the PBR. Nonetheless, the number of microalgae between

the two samples seems to be very close after stabilization is achieved. This data shows

that the microalgae growth is quite low in the inner locations within the PBR due to a weak

access to light.

The major increase of the heterotrophic bacteria was on day 27 which coincides

with the day of the major decrease on the peripheral plates. However no relation between

the two events could be found. After this day, the heterotrophic bacteria decreased

reaching very low values on day 41. As for the peripheral samples, this decrease may be

explained by the peak of microalgae on the plates as already discussed. On day 48 a

recovery of the heterotrophic bacteria was observed reaching the 2.08x104 and 1.89x104

CFUs/cm2 on the 0.5C and 1.5C plates respectively.

The time gap between the first increase, decrease and second increase is the

same for all samples. There was a two week gap between the first increase and the

decrease, and a gap of one week between this last and the second increase. This pattern

is similar to all samples which may indicate that the same mechanisms may be acting on

all plates in spite of the different environmental conditions.

Further molecular analysis to the microbial community could provide information

on this dynamics and the identification of these bacteria would provide more clues about

the factors that initiate this process.

- Microalgae isolation

The dominant microalgae was successfully isolated (Figure 40) and was

morphologically identified as a Chlorella sp., confirming that this is the dominant algae

observed in the biofilm of the PBRs.

This microalga presents a very high resistance against the leachate contaminants

what makes it an interesting candidate for future applications. Further characterization

studies could indicate possible applications of this microalga and improve our knowledge

about its resistance and metabolism, allowing a better understanding of its role on the

leachate treatment which could lead to an optimization of the leachate treatment using this

organism.


60

Figure 40 – Microalgae isolated from the phototrophic biofilm.

Conclusions

61

Conclusions

Conclusions

62

Conclusions

The treatment configuration applied on this pilot treatment plant showed low

decontamination efficiency which could be improved if treatment steps are optimized and

these are pointed out below.

The O3 treatment needs a bigger O3 mixer to improve the hydraulic retention time

and consequently the reaction time between the O3 and the leachate. The big

recirculation should be stopped because O3 is acting on the suspended cells coming from

the biological treatment, what reduces the amount of O3 that reacts with the refractory

organic matter present in the untreated leachate.

All biological treatments presented a rather low efficiency. In the trickling filter this

is explained by the low biodegradability of the leachate, the high concentrations of NH4-N

and accumulation of NO2-N within the system. These last two also affect the PBRs

function, due to microalgae inhibition. In the PBRs some design problems were also

observed. The biofilm formed on the wall prevents light from reaching the inner regions of

the PBRs where a larger amount of available surface for biofilm formation exists. Future

designs should have a wall cleaning system to remove this biofilm. As the amount of CO2

could be a limiting factor for the microalgae growth, CO2 should be added to the system in

order to increase microalgae growth or the leachate flux may be increased, also

increasing the CO2 supply.

Improving the leachate retention time of the whole system, namely the trickling

filter and clarifier, would also allow better removal rates. Further tests should be

performed to assess this and allow the comparison between efficiency and cost.

The biofilm formed inside the PBRs differs mainly between the inner and

peripheral zones due to light access. The formed peripheral biofilm was in its majority a

thick phototrophic biofilm that absorbed light. The inner biofilm development had a curious

dynamics which alternated between pigmented and less pigmented biofilm. This had the

capacity of reflecting light, which could be used as an advantage in future PBR designs.

The PBRs different diameters did not reveal significant treatment efficiency differences

between them, possibly because of the low efficiency revealed by all PBRs. Further

treatment improvements, as the increase of the whole system hydraulic retention time,

may reveal differences in treatment between the different diameters.

A more precise characterization of the biofilm community could provide information

about the dynamics observed, permit a better comprehension of the whole system and

allow introducing improvements in the system to achieve a better efficiency.

Conclusions

63

The dominant organism present in the PBRs was a Chlorella sp. that has been

isolated for further characterization due to its resistance to a highly contaminated

environment as this landfill leachate. It could have future applications for wastewater

treatment or compounds production.

References

64

References

Amokrane A, Comel C, Veron J. 1997. Landfill leachates pretreatment by coagulation-

flocculation. Water Res. 31: 2775-2782.

Bai T, Lei H, Yu G, Yu Q, Li Z, Li H. 2009. High nitrite accumulation and strengthening

denitrification for old-age landfill leachate treatment using an autocontrol two-stage

hybrid process. Process Saf. Environ. Protection. 87:307-314.

Bashir MJK, Isa MH, Kutty SRM, Awang ZB, Aziz HA, Mohajeri S, Farooqi IH. 2009.

Landfill leachate treatment by electrochemical oxidation. Waste Manag. 29: 2534-2541.

Baun A, Ledin A, Reitzel LA, Bjerg PL, Christensen TH. 2004. Xenobiotic organic

compounds in leachates from ten Danish MSW landfills--chemical analysis and toxicity

tests. Water Res. 38: 3845-3858.

Becker EW. 1988. Micro-algae for human and animal consumption. In: Micro-algal

biotechnology. Borowitzka MA, Borowitzka LJ, ed. Cambridge University press.

Cambridge. p. 222–256.

Bernard C, Guido P, Colin J, Le Dû-Delepierre A. 1996. Estimation of the hazard of

landfills through toxicity testing of leachates--I. Determination of leachate toxicity with a

battery of acute tests. Chemosphere. 33: 2303-2320.

Bila DM, Filipe Montalvão A, Silva AC, Dezotti M. 2005. Ozonation of a landfill leachate:

evaluation of toxicity removal and biodegradability improvement. J. Hazard. Mater. 117:

235-242.

Borde X, Guieysse B, Delgado O, Muñoz R, Hatti-Kaul R, Nugier-Chauvin C, Patin H,

Mattiasson B. 2003. Synergistic relationships in algal-bacterial microcosms for the

treatment of aromatic pollutants. Bioresourc. Technol. 86: 293-300.

Borowitzka MA. 1998. Limits to growth, in Wastewater treatment with algae, Wong YS,

Tam NFY, ed. Springer Verlag. p. 203–226.

Brennan L, Owende P. 2010. Biofuels from microalgae--A review of technologies for

production, processing, and extractions of biofuels and co-products. Renew. Sust.

Energ. Rev. 14: 557-577.

Chan YJ, Chong MF, Law CL, Hassell DG. 2009. A review on anaerobic-aerobic

treatment of industrial and municipal wastewater. Chem. Eng. J. 155: 1-18.

Chavan A, Mukherji S. 2008. Treatment of hydrocarbon-rich wastewater using oil

degrading bacteria and phototrophic microorganisms in rotating biological contactor:

Effect of N:P ratio. J. Hazard. Mater. 154: 63-72.

Dave G, Nilsson E. 2005. Increased reproductive toxicity of landfill leachate after

degradation was caused by nitrite. Aquatic Toxicol. 73: 11-30.

References

65

de-Bashan LE, Moreno M, Hernandez J-P, Bashan Y. 2002. Removal of ammonium and

phosphorus ions from synthetic wastewater by the microalgae Chlorella vulgaris

coimmobilized in alginate beads with the microalgae growth-promoting bacterium

Azospirillum brasilense. Water Res. 36: 2941-2948.

de-Bashan LE, Hernandez J-P, Morey T, Bashan Y. 2004. Microalgae growth-promoting

bacteria as "helpers" for microalgae: a novel approach for removing ammonium and

phosphorus from municipal wastewater. Water Res. 38: 466-474.

de Godos I, Vargas VA, Blanco S, González MCG, Soto R, García-Encina PA, Becares E,

Muñoza R. 2010. A comparative evaluation of microalgae for the degradation of

piggery wastewater under photosynthetic oxygenation. Bioresourc. technol. 101:5150-

5158.

EEA. 2001. Indicator Fact Sheet Signals 2001 – Chapter Waste. In Agency EE, ed. Total

waste generation. Luxembourg. p.6.

EPA. 2009. Municipal solid waste generation, recycling, and disposal in the United States,

detailed tables and figures for 2008. U.S. Environmental Protection Agency - Office of

Resource Conservation and Recovery, ed. p.58.

Eurostat. 2005. Waste generated and treated in Europe. Office for official publications of

the European communities, ed. Luxembourg.

Fallowfield HJ, Cromar NJ, Evison LM. 1996. Coliform die-off rate constants in a high rate

algal pond and the effect of operational and environmental variables. Water Sci

Technol. 34: 141-147.

Fierro S, Saavedra MPS, Copalcúa C. 2008. Nitrate and phosphate removal by chitosan

immobilized Scenedesmus. Bioresourc. Technol. 99:1274-1279.

Fukami K, Nishijima T, Ishida Y. 1997. Stimulative and inhibitory effects of bacteria on the

growth of microalgae. Hydrobiologia. 358:185-191.

Gonzalez L, Cañizares RO, Baena S. 1997. Efficiency of ammonia and phosphorus

removal from a colombian agroindustrial wastewater by the microalgae Chlorella

vulgaris and Scenedesmus dimorphus. Bioresourc. Technol. 60:259-262.

Greenberg AE, Clesceri LS, Eaton AD. 1992. Standard methods for the examination of

water and wastewater. 18th edition.

Grönlund E. 2002. Microalgae at wastewater treatment in cold climate. Graduation thesis.

Lulea University of Technology. Lulea. p. 6.

Guzzon A, Bohn A, Diociaiuti M, Albertano P. 2008. Cultured phototrophic biofilms for

phosphorus removal in wastewater treatment. Water Res. 42: 4357-4367.

Huo S, Xi B, Yu H, He L, Fan S, Liu H. 2008. Characteristics of dissolved organic matter

(DOM) in leachate with different landfill ages. J. Environ. Sci. 20: 492-498.

References

66

Jiménez-Pérez MV, Sánchez-Castillo P, Romera O, Fernández-Moreno D, Pérez-

Martínez C. 2004. Growth and nutrient removal in free and immobilized planktonic

green algae isolated from pig manure. Enzyme Microb. Technol. 34: 392-398.

Kulikowska D, Klimiuk E. 2008. The effect of landfill age on municipal leachate

composition. Bioresourc. Technol. 99: 5981-5985.

Kurniawan TA, Lo W, Chan GYS. 2006. Physico-chemical treatments for removal of

recalcitrant contaminants from landfill leachate. J. Hazard. Mater. 129: 80-100.

Larsdotter K. 2006. Wastewater treatment with microalgae – a literature review. VATTEN

62:31-38.

Ledakowicz S, Kaczorek K. 2004. Laboratory Simulation of Anaerobic Digestion of

Municipal Solid Waste. Journal of Environmental Science and Health, Part A: Tox.

Hazard. Subst. Environ. Eng. 39: 859 - 871.

Lee K, Lee C. 2002. Nitrogen removal from wastewaters by microalgae without

consuming organic carbon sources. J. Microbiol. Biotechnol. 12:979-985.

Lin L, Chan GYS, Jiang BL, Lan CY. 2007. Use of ammoniacal nitrogen tolerant

microalgae in landfill leachate treatment. Waste Manag. 27: 1376-1382.

Linde K, Jönsson A, Wimmerstedt R. 1995. Treatment of three types of landfill leachate

with reverse osmosis. Desalination 101: 21-30.

Lopez A, Pagano M, Volpe A, Claudio Di Pinto A. 2004. Fenton's pre-treatment of mature

landfill leachate. Chemosphere 54: 1005-1010.

Kim D, Lee D, Keller J. 2006. Effect of temperature and free ammonia on nitrification and

nitrite accumulation in landfill leachate and analysis of its nitrifying bacterial community

by FISH. Bioresourc. technol. 97:459-68.

Mata TM, Martins AA, Caetano NS. 2010. Microalgae for biodiesel production and other

applications: A review. Renew. Sust. Energ. Rev. 14: 217-232.

Matthews R, Winson M, Scullion J. 2009. Treating landfill leachate using passive aeration

trickling filters; effects of leachate characteristics and temperature on rates and process

dynamics. Sci. Total Environ. 407: 2557-2564.

Molinuevo-Salces B, García-González MC, González-Fernández C. 2010. Performance

comparison of two photobioreactors configurations (open and closed to the

atmosphere) treating anaerobically degraded swine slurry. Bioresourc. Technol. 101:

5144-5149.

Moreno-Garrido I. 2008. Microalgae immobilization: Current techniques and uses.

Bioresourc. Technol. 99: 3949-3964.

Mühling M, Allen JW, Murrell JC, Joint I. 2008. Improved group-specific PCR primers for

denaturing gradient gel electrophoresis analysis of the genetic diversity of complex

microbial communities. ISME J 2:379-392.

References

67

Muñoz R, Guieysse B. 2006. Algal-bacterial processes for the treatment of hazardous

contaminants: A review. Water Res. 40: 2799-2815.

Muñoz R, Köllner C, Guieysse B. 2009. Biofilm photobioreactors for the treatment of

industrial wastewaters. J. Hazard. Mater. 161: 29-34.

Muñoz R, Köllner C, Guieysse B, Mattiasson B. 2004. Photosynthetically oxygenated

salicylate biodegradation in a continuous stirred tank photobioreactor. Biotechnol.

Bioeng. 87: 797-803.

Oilgae. 2009.Oilgae Guide to Algae-based Wastewater Treatment. Oilgae. India.

Oron G, Shelef G, Levi A, Meydan A, Azov Y. 1979. Algae/bacteria ratio in high-rate

ponds used for waste treatment. Appl. environ. microbial. 38:570-576.

Oswald WJ, 1988. Micro-algae and waste-water treatment. In: Borowitzka MA, Borowitzka

LJ, ed. Micro-algal Biotechnology. Cambridge University Press, Cambridge, pp. 305–

328.

Ozturk I, Altinbas M, Koyuncu I, Arikan O, Gomec-Yangin C. 2003. Advanced physico-

chemical treatment experiences on young municipal landfill leachates. Waste Manag.

23: 441-446.

Pambrun V, Paul E, Sperandio M. 2008. Control and modelling of partial nitrification of

effluents with high ammonia concentrations in sequencing batch reactor. Chem. Eng.

Processing: Process Intensification. 47:323-329.

Pulz, Pulz O. 2001. Photobioreactors: production systems for phototrophic

microorganisms. Appl. Microbiol. Biotechnol. 57: 287-293.

Renou S, Givaudan JG, Poulain S, Dirassouyan F, Moulin P. 2008. Landfill leachate

treatment: Review and opportunity. J. Hazard. Mater. 150: 468-493.

Roeselers G, van Loosdrecht M, Muyzer G. 2007. Heterotrophic Pioneers Facilitate

Phototrophic Biofilm Development. Microb. Ecol. 54: 578-585.

Safonova E, Kvitko KV, Iankevitch MI, Surgko LF, Afti IA, Reisser W. 2004. Biotreatment

of Industrial Wastewater by Selected Algal-Bacterial Consortia. Eng. Life Sci. 4: 347-

353.

San I, Onay TT. 2001. Impact of various leachate recirculation regimes on municipal solid

waste degradation. J. Hazard. Mater. 87: 259-271.

Sang N, Li G. 2004. Genotoxicity of municipal landfill leachate on root tips of Vicia faba.

Mutat Res. 560: 159-165.

Schumacher G, Blume T, Sekoulov I. 2003. Bacteria reduction and nutrient removal in

small wastewater treatment plants by an algal biofilm. Water sci. technol. 195-202.

Schwarzbauer J, Heim S, Brinker S, Littke R. 2002. Occurrence and alteration of organic

contaminants in seepage and leakage water from a waste deposit landfill. Water Res.

36: 2275-2287.

References

68

Silva AC, Dezotti M, Sant'Anna GL. 2004. Treatment and detoxification of a sanitary

landfill leachate. Chemosphere. 55: 207-214.

Spolaore P, Joannis-Cassan C, Duran E, Isambert A. 2006. Commercial applications of

microalgae. J. Biosci. Bioeng. 101: 87-96.

Tsilogeorgis J, Zouboulis A, Samaras P, Zamboulis D. 2008. Application of a membrane

sequencing batch reactor for landfill leachate treatment. Desalination. 221: 483-493.

Visvanathan C, Choudhary MK, Montalbo MT, Jegatheesan V. 2007. Landfill leachate

treatment using thermophilic membrane bioreactor. Desalination. 204: 8-16.

Wang LK, Wu Z, Shammas NK. 2009. Trickling filters. In: Wang LK, Pereira NC, Hung YT,

Shammas NK, ed. 2009. Handbook of environmental engineering, volume 8: biological

treatment processes. Humana Press. p. 371-434.

Wantawin C, Juateea J, Noophan PL, Munakata J. 2008. Autotrophic nitrogen removal in

sequencing batch biofilm reactors at different oxygen supply modes. Water sci. technol.

58:1889-94.

Wiszniowski J, Robert D, Surmacz-Gorska J, Miksch K, Weber J. 2006. Landfill leachate

treatment methods: A review. Environ. Chem. Lett. 4: 51-61.

Wolfaardt GM, Lawrence JR, Robarts RD, Caldwell DE. 1994. The role of interactions,

sessile growth, and nutrient amendments on the degradative efficiency of amicrobial

consortium. Can. J. Microbiol. 40:331-340.

Wu JJ, Wu C-C, Ma H-W, Chang C-C. 2004. Treatment of landfill leachate by ozone-

based advanced oxidation processes. Chemosphere. 54: 997-1003.

Xu ZY, Zeng GM, Yang ZH, Xiao Y, Cao M, Sun HS, Ji LL, Chen Y. 2010. Biological

treatment of landfill leachate with the integration of partial nitrification, anaerobic

ammonium oxidation and heterotrophic denitrification. Bioresourc. technol. 101:79-86.

landfill leachate treatment: a new photobioreactor technology mct… · table 2 - advantages and...

Documents