Nitrification in Saline Industrial Wastewater

NITRIFICATION IN SALINE INDUSTRIAL WASTEWATER Nitrification in Saline Industrial Wastewater DISSERTATION Submitted in...

Author: Moustafa Samir Moussa

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NITRIFICATION IN SALINE INDUSTRIAL WASTEWATER

Nitrification in Saline Industrial Wastewater DISSERTATION Submitted in fulfilment of the requirements of the Board for Doctorates of Delft University of Technology and of the Academic Board of the UNESCO-IHE Institute for Water Education for the Degree of DOCTOR to be defended in public on Monday, 29 March 2004 at 10:30 hours in Delft, The Netherlands by

MOUSTAFA SAMIR MOUSSA born in Cairo, Egypt Master of Science, UNESCO-IHE

This dissertation has been approved by the promotor Prof.dr. H.J.Gijzen Prof.dr.ir. M.C.M.van Loosdrecht Members of the Awarding Committee: Chairman

Rector Magnificus Delft University of Technology

Co-chairman

Director UNESCO-IHE, Delft

Prof.dr. H.J.Gijzen

UNESCO-IHE, Delft, promotor

Prof.dr.ir. M.C.M.van Loosdrecht

Delft University of Technology, promotor

Prof.dr. J.G.Kuenen

Delft University of Technology

Prof.dr. P.Wilderer

Technical University München, Germany

Dr.ir. A.Klapwijk

Wageningen University

Dr. H.J.Lubberding

UNESCO-IHE, Delft

This research was sponsored by BTS Senter (BTS99130), Shell Global Solutions International, The Hague, Heiploeg Shrimp Processing, Zoutkamp and Ecco Tannery, Dongen. The project was carried out at the departments of Environmental Resources, (UNESCO-IHE, Delft) and of Biotechnology (Delft University of Technology). Copyright © 2004 Taylor & Francis Group plc, London, UK All rights reserved No part of this publication or the information contained herein may be reproduced, stored in a retrieval system, or transmitted in any form or by any means, electronic, mechanical, by photocopying, recording or otherwise, without written prior permission from the publisher. Although all care is taken to ensure the integrity and quality of this publication and the informationherein, no responsibility is assumed by the publishers nor the authors for any damage to property or persons as a result of operation or use of this publication and/or the information contained herein. Published by A.A.Balkema Publishers, a member of Taylor & Francis Group plc. http://www.balkema.nl/ and http://www.tandf.co.uk/ This edition published in the Taylor & Francis e-Library, 2006. “To purchase your own copy of this or any of Taylor & Francis or Routledge’s collection of thousands of eBooks please go to http://www.ebookstore.tandf.co.uk/.” ISBN 0-203-02454-0 Master e-book ISBN

ISBN 90 5809 671 8 (Print Edition) (A.A.Balkema Publishers)

Contents Symboles

vii

Summary

x

Chapter1 Introduction Chapter2 Improved method for determination of ammonia and nitrite oxidation activities in mixed bacterial cultures Chapter3 Short term effects of various salts on ammonia and nitrite oxidisers in enriched bacterial cultures Chapter4 Long Term Effects of Salt on Activity, Population Structure and Floc Characteristics in Enriched Bacterial Cultures of Nitrifiers Chapter5 Modelling Nitrification, Heterotrophic growth and Predation in Activated Sludge Chapter6 Nitrification activities in full-scale treatment plants with varying salt loads Chapter7 Model-based evaluation of the upgrading of a full-scale industrial wastewater treatment plant Chapter8 Evaluation and Outlook

4 32 48 69 95 124 140 156

Samenvatting (Summary in Dutch)

160

Acknowledgements

162

Curriculum

164

Symbols ASM

Activated Sludge Model

BOM

Biological Oxygen Monitor

BR

Batch Reactor

bNH4

Aerobic decay rate of ammonia oxidisers (day−1)

bNO2

Aerobic decay rate of nitrite oxidisers (day−1)

bH

Aerobic decay rate of heterotrophic biomass (day−1)

bpredators

Aerobic decay rate of predators (day−1)

Ci

Concentration of ionic species, i (mole)

Ci−

Concentration of anion species, i− (mole)

Ci+

Concentration of cation species, i+ (mole)

COD

Chemical Oxygen Demand

DO

Dissolved Oxygen

Fxi

Fraction of inert COD generated in biomass lysis

FISH

Fluorescent In Situ Hybridisation

HRT

Hydraulic Retention Time (day)

iN,XI

Nitrogen Content of XI (g N/g COD)

iN,BM

Nitrogen Content of biomass (g N/g COD) Affinity constant for ammonia of ammonia oxidisers (mg N/L) Affinity constant for oxygen of ammonia oxidisers (mg O2/L) Affinity constant for nitrite of nitrite oxidisers (mg N/L) Affinity constant for oxygen of nitrite oxidisers (mg O2/L) Affinity constant for nitrate of heterotrophic biomass (mg N/L) Affinity constant for organic carbon of heterotrophic biomass (mg COD/L) Affinity constant for oxygen of heterotrophic biomass (mg O2/L) Affinity constant for oxygen of predators (mg O2/L)

MLSS

Mixed Liquor Suspended Solids (mg/L)

MLVSS

Mixed Liquor Volatile Suspended Solids (mg/L)

mNH4

maintenance coefficient of ammonia oxidisers (mg NH4–

N/gXN– COD.day) mNO2

maintenance coefficient of nitrite oxidisers (mg NO2– N/gXN– COD.day)

mH

maintenance coefficient of heterotrophic biomass (mg COD/gXH– COD.day)

MNH4

Monod term for ammonia in bacterial growth

OUR

Oxygen Uptake Rate (mg O2/L.h)

pKa

The negative logarithm of stoichiometric dissocial constant

RNH4

volumetric uptake rate of ammonia (mg NH4–N/L.h)

RNO2

volumetric uptake rate of nitrite (mg NO2–N/L.h)

Rmax,NH4

maximum volumetric uptake rate of ammonia (mg NH4– N/L.h)

Rmax,NO2

maximum volumetric uptake rate of nitrite (mg NO2– N/L.h)

SBR(s)

Sequencing Batch Reactor (s)

SRT

Sludge Retention Time

SO2

Concentration of Oxygen (mg O2/L)

SNH4

Concentration of ammonia (mg NH4–N/L)

SNO3

Concentration of nitrate (mg NO3–N/L)

SNO2

Concentration of nitrite (mg NO2–N/L)

SN2

Concentration of nitrogen (mg N/L)

SS

Concentration of organic substrate (mg COD/L)

T

The temperature in °K

VFA

Volatile Fatty Acids

VSS

Volatile Suspended Solids

WWTP(s)

Wastewater Treatment Plant (s)

XNH4

Concentration of ammonia oxidisers (mg-VSS/L)

XH

Concentration of heterotrophic biomass (mg-VSS/L)

Xpredators

Concentration of predators (mg-VSS/L)

XI

Concentration of particulate inert (mg-VSS/L)

YNH4

Yield coefficient of ammonia oxidisers per NO2–N (g COD/g NH4–N)

YNO2

Yield coefficient of nitrite oxidisers per NO3−–N (g COD/g NO2–N)

YH

Yield coefficient of heterotrophic biomass on SS (g COD/g COD)

Ypred

Yield coefficient of predators on bacteria (g COD/g COD)

Zi

Charge of species, i

Zi−

Charge of anion species, i−

Zi+

Charge of cation species, i+

ηNH4

Anoxic reduction factor for ammonia oxidisers decay

ηNO2

Anoxic reduction factor for nitrite oxidisers decay

ηH

Anoxic reduction factor for heterotrophic growth Maximum growth rate of ammonia oxidisers (day−1) Maximum growth rate of nitrite oxidisers (day−1) Maximum growth rate of heterotrophic biomass (day−1)

µpredators

Growth of predators, presented in the model as predation rate (day−1)

Summary Biological nitrification-denitrification is one of the most common processes for nitrogen removal from wastewater. However, nitrification, the rate limiting step in biological nitrogen removal, proved to be one of the most difficult processes to design and control in wastewater treatment plants, because nitrifying bacteria are slow-growing and very sensitive to environmental factors (temperature, pH, dissolved oxygen concentration, toxic and inhibitory compounds). Researchers so far have concentrated mainly on nitrification in domestic wastewater treatment and achieved broad knowledge and practical experience about the process. The result is that biological nitrogen removal is widely and successfully applied for municipal wastewater. However, these experiences are not directly applicable to industrial wastewater due to its specific composition. Several industries are dealing with a high salt concentration in their wastewater. Also the policy of more economic use of water and water reuse will result in an increase of salt content of the ultimately produced wastewater. High salt levels may negatively affect nitrification, demonstrating the need for improved understanding of the precise effects of salt on nitrification. The response of the nitrification process to saline conditions and the adaptation mechanisms of nitrifying bacteria towards these conditions are still unknown. The available studies on the effect of salt on nitrification show a decline in activity for ammonia and nitrite oxidisers. However, it does not give clear answers on: what are the main inhibiting factors causing the effects, do all salts have similar effects, what is the maximum acceptable salt level, are ammonia oxidisers or nitrite oxidisers most sensitive to salt stress, can nitrifiers adapt to long term salt stress and are some specific nitrifiers more resistant to salt stress than others? The main focus of this dissertation is the understanding of the effects of salinity on nitrification considering all these questions. The research was carried out in two phases. In the first phase, laboratory scale activities were conducted to obtain fundamental data to determine the relationship between salinity and nitrification. In the second phase the results collected from the laboratory experiments were compared and validated with the results collected from full-scale treatment plants. Modelling was employed in both phases to provide a mathematical description for salt inhibition on nitrification and to facilitate the comparison. First phase: A method to measure the activity of ammonia and nitrite oxidisers in mixed bacterial cultures was developed and applied in the research as standard method to determine the inhibition effects of salt on ammonia and nitrite oxidisers. The short-term effects of various types of salt on the activity of ammonia and nitrite oxidisers were studied. Different types of salts appeared to have different inhibition effects on the ammonia and nitrite oxidisers. Non-adapted and adapted (to 10 g NaCl–Cl/L for one year) enriched cultures of nitrifiers were used to investigate the long-term effect of salt (gradually increased with 5 g Cl−/L up to 40 g Cl−/L). No difference in activity was observed between the adapted and non-adapted sludge. At 40 g Cl−/L inhibition reached

95% of salt free activity for ammonia and nitrite oxidisers in both adapted and nonadapted reactors. Nitrosomonas europaea and Nitrobacter sp were the only nitrifiers present at high salt levels. Increased salt concentrations resulted in better settling characteristics of the nitrifying sludge. At the same time the protozoan and metazoan predators in the laboratory scale experiments were found to be affected by salt. This effect was used to develop a mathematical model to describe the interaction between nitrifiers, heterotrophs and predators. Second phase: Nitrifier activities and population structure in full-scale domestic and industrial wastewater treatment plants (WWTPs) operated under various salt levels were investigated and compared with results obtained from laboratory scale activities. Finally, the activated sludge model No. 1 (ASM 1) was modified and applied to simulate COD and nitrogen removal in a full-scale industrial WWTP operated under salt stress. The research has lead to an improved understanding of the effect of salinity on nitrification. The results obtained within the course of this research can be used to improve the sustainability of the existing WWTPs operated under salt stress. The findings also form a guideline for more economical and sustainable design and start up of WWTPs dealing with salt in future.

Chapter 1 Introduction

Abstract The global situation for nitrogen is getting out of hand. There is a serious imbalance between the influx and efflux of N in the biosphere. The direct cause is the rapidly increasing production of chemical fertilisers. The annual production of fertiliser N has increased 9 fold over the past 40 years and amounts currently to some 37% of the world-wide biological Nfixation. Such a massive introduction of reactive forms of nitrogen into the environment over a relatively short period of time has numerous deleterious consequences, causing environmental and public health problems, both locally and at a global scale. The response to increasing pollution problems necessitated the promulgation of effluent standards for nutrients. In this framework environmental legislation in most countries includes stringent limitations for nitrogen to be discharged. However, the implementation of effluent standards at a global scale is limited due to the phenomenal costs of the high-rate wastewater treatment technology. It remains a challenge to come up with nitrogen pollution control strategies, which are effective and low cost. Other sources of nitrogen pollution than domestic should be considered. Industrial wastewater not only represents twice the volume of domestic wastewater, but also is usually more concentrated. Having a cost-effective N-removal technology in industry is still a target and needs more attention. Biological nitrificationdenitrification is the most common processes for nitrogen removal from wastewater; nitrification is the rate-limiting step in biological nitrogen removal. Nitrification in industrial wastewater presents a number of difficulties, including a wide range of different and varying temperatures, pH, presence of toxic compounds and salinity. Studies on the effect of salt on nitrification show a decline in activity for ammonia and nitrite oxidisers. However, no information is available on the maximum acceptable salt level and which nitrifying group is most sensitive to salt stress. The need for understanding the precise effects of salt on nitrification was addressed, as the main aim of this study.

1 Introduction 1.1 The nitrogen cycle Nitrogen is an essential component to all living organisms, as it is an important atom of DNA, RNA, proteins and other key organic molecules. In general, living organisms contain between 10–15% of their biomass as nitrogen. Although N represents only a minor constituent of living matter, it has been and continues to be the main limiting factor for biomass production on a global scale. Also in agricultural production, it appears that the other two limiting nutrients, potassium and phosphorus, are less frequently the prime limiting factor (Smil, 1997). Nitrogen is present on earth in many forms and huge amounts are stored in sediment and rock deposits and in the atmosphere. Nitrogen is present in a variety of compounds with different oxidation states. The movement and transformation of these nitrogen compounds through the biosphere is characterised by the nitrogen cycle (Figure 1.1). The atmosphere serves as a reservoir of nitrogen in the form of nitrogen gas, which makes up about 78% of the atmosphere, but nitrogen in this form is too inert to play a direct role in ecosystems. Plants and animals cannot use nitrogen gas directly from the air as they do with carbon dioxide and oxygen. It is only accessible to N2—fixing bacteria. The nitrogen must be available in a reactive form with hydrogen or oxygen before it can be assimilated by plants or used by other organisms. The plants, in turn, can be consumed by animals for the generation of animal protein.

Figure 1.1 The nitrogen cycle Transformation of these nitrogen compounds can occur through several mechanisms. Those of importance include N-fixation, ammonification, synthesis, nitrification, and denitrification. Each can be carried out by particular microorganisms.

Introduction

5

Nitrogen fixation Fixation of nitrogen (physical, chemical or biological) means the incorporation of inert, gaseous nitrogen into chemical compounds that eventually can be used by living organisms. Biological fixation of N2 is prominently accomplished by specialized microorganisms: cyanobacteria, symbiotic and free-living bacteria. Lightning also indirectly transforms atmosphere nitrogen into nitrate, which rains onto soil. Finally, N2 can be fixed industrially by the Haber-Bosch process, invented in 1913. At present the industrial fixation of nitrogen into ammonia plays a significant role, because it is responsible for 30% of the total nitrogen influx into the biosphere (Gijzen and Mulder 2001). Ammonification In most ecosystems nitrogen is preliminary stored in living and dead organic mater. Ammonification is the process responsible for the change of organic nitrogen compounds into the ammonia form. In general, ammonification occurs during decomposition of animal and plant tissue and animal faecal matter by bacteria; after hydrolysis of the proteins, the amino acids are either reused or the amino groups are converted into ammonia. Also the nitrogen present in urine is—via urea—converted into ammonia. Nitrification Nitrification is the biological oxidation of ammonium. This is done in two steps, first to the nitrite form, then to the nitrate form. Both steps can be carried out by different genera, both using CO2 as their source of cellular carbon. These transformation reactions are generally coupled and proceed rapidly to the nitrate form; under normal conditions nitrite levels are usually very low. The produced nitrate is used either by plants in the assimilation process or reduced by denitrification to N2. Denitrification Denitrification is the biological reduction of nitrate to nitrogen gas. It can proceed through several steps in the biochemical pathway, with the ultimate production of nitrogen gas. A fairly broad range of heterotrophic bacteria is involved in the process, requiring an organic carbon source for energy (Kuenen and Robertson 1994; Schmidt et al 2003). Nitrate reduction to ammonia In contrast to denitrification, the process of dissimilatory nitrate reduction to ammonia (DNRA) does not have N2 but NH4+ as final product. Apart from a nitrate reductase, a nitrite reductase, which reduces nitrite to ammonia, is involved in this process. Denitrification and DNRA can occur simultaneously and DNRA can be of quantitative importance in environments with high carbon/nitrate ratio or high sulphide concentration (Brunet and Garcia-Gil 1996; Cole 1996; Simon 2002).

Nitrification in saline industrial wastewater

6

ANAMMOX The denitrifying bacteria (as described above) are not the only bacteria producing nitrogen gas. Ammonia can be oxidized under anaerobic conditions also leading to N2 and it became clear that slow growing autotrophic bacteria belonging to the order of the Planctomycetales are carrying out this process. This process, in which both ammonia and nitrite are converted to N2, is called ANAMMOX, an acronym for ANaerobic AMMonia OXidation (Mulder et al 1995; Schmidt et al 2003). Assimilation Assimilation is the process in autotrophic organisms in which nitrogen compounds (NH4+, NO3−) are incorporated into cell material for growth, a biochemical mechanism that uses ammonia or nitrate. Animals and other heterotrophic organisms require protein from plants and other animals as their nitrogen source. They are not capable of transforming inorganic nitrogen into an organic nitrogen form. 1.2 The nitrogen cycle out of balance The influx and efflux of N in the biosphere has been kept in balance by nature. Several decades ago this balanced situation started to undergo a radical change mainly due to binding of atmospheric nitrogen gas for the manufacturing fertilisers (the invention of ammonia synthesis by Fritz Haber). The first commercial ammonia factory started its operations in 1913 in Germany, but production levels at a global scale remained low until the process became more energy efficient due to technological innovations in the 1960s. Since then the production of industrial nitrogen fertiliser via the so-called Haber-Bosch process showed a sharp increase. This process has removed the fundamental restriction on food production and therefore on population growth. Indeed, the doubling of the world population over the last 40 years would not at all have been possible without the intensive agriculture and animal production systems which primarily depend on nitrogen fertiliser. The increase in production of nitrogen fertiliser has been much faster than population increase. While population doubled between 1960 and 2000, the annual production of fertiliser nitrogen increased nine-fold from 1×1010 to 9×1010 kg. Current production is equivalent to about 37% of the total amount of nitrogen input achieved via terrestrial and marine biological N2 fixation (about 24×1010 kg per year). There is probably no other elemental cycle where the human impact has been so dramatic as the case for nitrogen (Gijzen and Mulder 2001). The massive introduction of reactive forms of nitrogen into the environment over a relatively short period of time has numerous deleterious consequences, causing environmental and public health problems, both locally and at a global scale (Scheible and Heidman, 1994; Vitousek et al 1997; Wiesmann 1994): • The formation of blooms of toxic cyanobacteria in fresh waters is of considerable concern with respect to human and animal health (e.g. potable water supply, fish production). Eventually the produced cyanobacteria, algal and plant biomass will die

Introduction

7

and become subject to biodegradation, causing substantial oxygen depletion and biodiversity loss in water bodies. • The oxidation of ammonia released into the environment, either directly (eutrophication) or from biomass degradation, will result in low oxygen levels of affected water bodies (theoretical consumption 4.57 g O2/g−N). The released toxins from cyanobacteria and the lower levels of dissolved oxygen will obviously affect many species of aquatic life. • The lower oxygen levels in water bodies may also result in incomplete nitrificationdenitrification and therefore stimulate the formation of NO and N2O gasses. Together with increased methane production from the decomposition of plant biomass in the sediments, this could contribute significantly to the global greenhouse effect. • The effect of high levels of nitrite and nitrate in drinking water may cause the so-called ‘blue baby’ disease (methemoglobinemia) in infants. The relation between nitrates and some forms of cancer has also been reported. Despite these negative impact of the fertiliser use there is, at present, no other substitute available and therefore mankind will develop an increasing dependency on the HaberBosch synthesis while population grows over the coming decades. There is some hope that the nitrogen fixing capacities of Rhizobium could be incorporated directly into plant species via genetic engineering, but it seems realistic to assume that this ‘solution’ is still several decades away from becoming reality. Therefore, reliance on chemical fertiliser must further increase to provide sufficient protein for the growth of the additional 2 to 3 billion people that will be born during the next 50 years (Gijzen and Mulder 2001). At the same time however, finding a sollution for the removal of reactive forms of nitrogen from the envrionment poses another urgent challenge. 1.3 Actions for rebalancing the nitrogen cycle When analysing the nitrogen pollution, care must be taken that the source of pollution is identified, in order to take measures to solve the problem. Different sources can contribute to the nitrogen problem, differing from location to location. The challenge is to find out the most effective approach to tackle the most urgent and most important one. Moreover, the approach for the same problem in one specific location may not be applicable to another. 1.3.1 Sources of nitrogen pollution The sources of the nitrogenous compounds in water can be of human, industrial, as well as agricultural origin. Natural sources can be atmospheric precipitation, dust, non-urban and non-agricultural leachates and biological fixation. Nitrogenous compounds of human origin can be for example treated and non-treated sewage, agricultural leachates (for instance from excess addition of fertilisers), some industrial wastewaters and surface runoff. Sewage always contains nitrogenous compounds, typically in concentrations between 25–50 mg N/L. The nitrogen consists of approximately 60% ammonia nitrogen, 40% organically bound nitrogen, and a small amount of nitrate (Scheible and Heidman, 1994). When sludge from sewage treatment plants is digested, a rich nitrogenous flow is


8

produced, with ammonia concentrations of 700 to 1000 mg-N/L. Agricultural leaching can result in large amounts (5–25 kg per hectare) of nitrogenous compounds released per year (Scheible and Heidman, 1994). In industrial effluents ammonia concentrations are often much higher (Wiesmann, 1994); the ammonia and nitrate levels in some industrial wastewaters is given in Table 1.1.

Table 1.1 Ammonia and nitrate concentrations in industrial wastewaters (Wiesmann, 1994). Industry/Products sludge digestion

Ammonia concentration range (g N/L) 1

tannery

0.35

cokery

0.45–4.1

oil refinery

0.02–0.9

coal gasification

1–2.5

fertiliser

0.2–1

synthetic fibre

0.8

slaughterhouse

0.15

livestock: swine

2.3

livestock: cattle

0.5–2.3

rendering plant

0.8

dairy

0.6

distillery

1.5

cellulose and paper

0.25

Pharmaceuticals

1–6

0.1–0.4

explosives

glass

Nitrate concentration range (g N/L)

2–12.5

0.3–0.65 0.48

electronics

0.5–2

pectin

1–2.7

uranium processing

4–11.3

Not only the nitrogen concentration, but also the amount of water polluted with nitrogen compounds is playing a role. The quantities of the three main sources of pollution (agriculture, industry and municipalities) are presented in Table 1.2. It illustrates that 70% of the total water use is for agriculture, 20% for industry and only 10% for domestic purposes. Industries generate more wastewater with higher quantities of nitrogen than municipalities.

Introduction

9

Table 1.2 Global water use in the 20th century (Cosgrove and Rijsberman, 2000). Use (Cubic kilometres) Agriculture

Industry

Municipalities

1900

1950

1995

Withdrawal

500

1,100

2,500

Consumption

300

700

1,750

Withdrawal

40

200

750

Consumption

5

20

80

Withdrawal

20

90

350

Consumption

5

115

50

0

10

200

Withdrawal

600

1,400

3,800

Consumption

300

750

2,100

Reservoir (evaporation) Total

1.3.2 Establishing nitrogen limits for emission The best way to prevent the release of nitrogenous compounds into surface waters is avoiding production of these substances. Integral process improvements are necessary to decrease the amounts of nitrogenous compounds produced. Furthermore, if possible the nitrogenous compounds should be recycled and not broken down into the constituting elements. However, end-of-pipe techniques (like wastewater treatment facilities) remain necessary to get rid of the nitrogenous compounds. Treatment objectives and priorities in industrialised countries have been gradually tightened over the past decades. This resulted in the so-called first, second and third generation of treatment plants (Table 1.3).

Table 1.3 The phased expansion and upgrading of wastewater treatment plants in industrialised countries to meet ever stricter effluent standards (WHO/UNEP, 1997). Decade Treatment objective

Treatment Operations included

1950– 60

Suspended/coarse solids removal

Primary

Screening, removal of grit, sedimentation

1970

Organic matter degradation

Secondary

Biological oxidation of organic matter

1980

Nutrient reduction (eutrophication)

Tertiary

Reduction of total N and total P

1990

Micro-pollutant removal

Advanced

Physicochemical removal of micropollutants


10

As a consequence, a number of treatment technologies, unit operations and processes have been developed to achieve the required treatment level (Table 1.4).

Table 1.4. Classification of common wastewater treatment processes according to their level of advancement (WHO/UNEP, 1997). Primary

Secondary

Tertiary

Advanced

Bar or bow screen

Activated sludge

Nitrification

Chemical treatment

Grit removal

Extended aeration

Denitrification

Reverse osmosis

Primary sedimentation

Aerated lagoon

Chemical precipitation

Electrodialysis

Comminution

Trickling filter

Disinfection

Carbon adsorption

Oil/fat removal

Rotating bio-discs

(Direct) filtration

Selective ion exchange

Flow equalisation

Anaerobic treatment/UASB

Chemical oxidation

Hyperfiltration

pH neutralisation

Anaerobic filter

Biological P removal

Oxidation

Imhoff tank

Stabilisation ponds

Constructed wetlands

Detoxification

Constructed wetlands

Aquaculture

Aquaculture

The fast population growth, urbanisation and industrialisation, all of which impose high demands on local water resource quality and quantity, while simultaneously generating pollution, which affects the very same water resource. The response to increasing pollution problems in receiving waters, and the growing concern about water quality protection, necessitated the promulgation of effluent standards for nutrients, especially for sensitive areas. In this framework environmental legislation in most countries includes stringent limitations for nitrogen to be discharged. National technology-based standards were established, moving all wastewater treatment facilities to secondary level at minimum. EU Policy on nutrients emissions Pollution and degradation of Europe’s waters as well as an increasing awareness by citizens and policy makers led to increased efforts to address water pollution. This resulted in a “second wave” of EU water legislation. Its first results were, in 1991, the adoption of • The Urban Wastewater Directive, addressing water pollution from all settlements except the small villages, as well as a range of industries with biodegradable wastewater; and, • The Nitrates Directive, addressing water pollution by nitrates from agriculture.

Introduction

11

Urban Wastewater Directive The urban wastewater directive has set ambitious objectives; • Wastewater collection and treatment for all settlements above 2000 population equivalents (p.e.) • Biological (secondary) treatment as a general rule, plus nutrients removal where the affected receiving waters show an elevated level of nitrates and/or eutrophication. The deadlines for achieving these objectives are 1998, 2000 and 2005, respectively (depending on the size of the discharge and the character of the receiving waters). Member states may choose between nutrient control by effluent concentration or by removal efficiency (Table 1.5). One or both nutrient parameters may apply depending on the local and regional situation.

Table 1.5 Discharge limits for treatment plants in EU. Total phosphorus

Effluent concentration

Minimum removal efficiency

2 mg/L for plants≤100,000 p.e.

80%

1 mg/L for plants>100,000 p.e. Total nitrogen

15 mg/L for plants≤100,000 p.e.

70–80%

10 mg/L for plants>100,000 p.e.

The Nitrate Directive The main objective of the nitrate directive is to reduce water pollution caused or induced by nitrates from agriculture and to prevent further nitrogen pollution. To ensure this objective, EU members have to identify waters (surface waters and ground waters) affected by nitrate pollution, and waters, which could be affected by nitrate pollution. The EU member states have the choice either to designate individual vulnerable zones in accordance with these criteria, or to apply the more stringent provisions of the Directive over all their territory (this option has been taken up by Denmark, Germany, Luxembourg, the Netherlands and Austria). Outside those affected areas (vulnerable zones) they have to promote codes of good agricultural practice on a voluntary basis (Blöch 2001). Implementation of EU policy on nutrient emissions A good practice of the EU policy implementation has been noticed in many member states. For instance, in the Netherlands, the water boards started in 1999 to take initiatives to benchmark their treatment performances. This shows the performance of the sewage treatment plants as a whole, with regard to the removal of oxygen-binding substances, phosphate and nitrogen. A treatment performance of 100 % means: 75% phosphate removal, 75% nitrogen removal and 90% removal of oxygen-binding substances. The treatment performance of the Dutch water boards has improved from 86% in 1996 to more than 91% in 2002. These water boards have to achieve 75% phosphate-removal and 75% nitrogen-removal by 2006. Phosphate removal of 79% was achieved in 2002.


12

Considering N-removal, in 2002 the removal efficiency averaged 69.4%. It is not certain, whether the 75% target for nitrogen removal will be achievable by 2006, taking the limited planned investments into consideration. In 2002, 40% of all sewage treatment plants failed at some time to meet the individual discharge requirements. Rapid improvement of the nitrogen removal in the next few years (before 2006) is a point of attention (Postma et al. 2003). Due to the successful implementation of the Directive in many EU states European waters have started to change. Reports by the European Environment Agency (1998, 1999a) clearly show improvements. The number of heavily polluted rivers has declined significantly, in particular as the pressure from organic matter and phosphates in urban wastewater has decreased. While progress has been made in many areas, others are still in a deplorable situation. After 25 years of European water legislation, Europe’s waters are in need of more protection, in need of increased efforts to get them clean or to keep them clean. This is a demand not only from the scientific community and other experts, but also to an ever-increasing extent from citizens and environmental organisations. 1.3.3 New Technologies National water quality goals have influenced the development of advanced treatment technologies especially in the area of nutrient control. With respect to nitrogen control, four recent developments have to be mentioned: The SHARON process SHARON (Hellinga et al 1998; van Loosdrecht and Jetten, 1998) is an acronym for Single reactor High activity Ammonia Removal Over Nitrite. In this process, a completely mixed reactor is operated at short residence time (1–1.5 days) and high temperature (30–40°C) leading to the selective wash out of nitrite oxidisers. This results in only partial oxidation of ammonia to nitrite and subsequently reduction of the latter to nitrogen gas in the denitrification process. This route is more favourable than the conventional route due to savings of 25% in oxygen supply and 40% reduction in COD demand. The ANAMMOX process Another recently discovered process allows nitrite reduction with ammonium as electron donor to nitrogen gas. This anaerobic ammonium oxidation (ANAMMOX, Mulder et al 1995; Schmidt et al 2003) process can be combined with partial nitrification (SHARON) leading to a direct net conversion of ammonium to N2 gas. Which makes complete autotrophic ammonia removal possible as a sustainable pathway of nitrogen removal from wastewater. The CANON process CANON is an acronym for Completely Autotrophic Nitrogen removal Over Nitrite (Strous et al 1997). This concept is the combination of partial nitrification and

Introduction

13

ANAMMOX in a single, aerated reactor. This process has been tested extensively on laboratory scale (Slikers et al 1998, 2003). Although ANAMMOX requires strict anoxic conditions, nitrifiers and ANAMMOX organisms are able to coexist under oxygenlimited conditions. Therefore, CANON would need process control to prevent nitrite build-up by oxygen excess under ammonia limitation (fluctuation of ammonia load). The OLAND process The OLAND process (oxygen-limited nitrification and denitrification) is described as a new process for one-step ammonium removal without addition of COD (Kuai and Verstraete 1998). Recently, it was confirmed that OLAND is based on the CANON concept. (Pynaert et al 2004) The formation of thick biomfilm could create a favourable condition for nitrifiers and ANAMMOX organisms to coexist even under normal oxygen conditions. A good overview of recent nitrogen removal technologies can be found in Schmidt et al (2003). 1.3.4 Future activities The implementation of effluent standards is so far limited to developed countries. In developing countries effluents remain largely untreated due to the phenomenal costs of sewerage systems and of high-rate wastewater treatment technology (Gijzen 2003). The challenge for these countries will be to come up with sewage management strategies, which are effective and low cost. The urgency of reconsidering the current practises (conventional technologies of domestic wastewater treatment) in the light of sustainability becomes evident. Implementation of cleaner production approaches is promising to achieve sustainable urban water and nutrient management. This approach incorporates pollution prevention or minimisation, treatment for reuse and stimulation of natural self-purification capacity of the receiving environment. The implementation of this approach is challenging (Gijzen 2003; Gijzen and Bijlsma 2000), but governments should put more efforts to bring it into practice. Nutrients from other sources than domestic are still a major challenge in nitrogen pollution control. Industrial wastewater is not only the double amount of domestic wastewater, but also in most cases it is much more concentrated (Table 1.1, 1.2). Concurrently, industrial activity is growing very fast to meet the human demands. Coming up with a cost-effective N-removal technology in industry is still a challenge and needs more attention. The main difficulties of industrial wastewaters are temperature, pH, presence of toxic compounds, salinity and fluctuations in flow and composition. This study focuses on nitrogen removal in industrial wastewaters emphasising on nitrification under salt stress, which is a common destabilising factor in industrial wastewaters. Full understanding of nitrification in saline wastewater will lead to a costeffective technology for industries reducing the pollution problems in the receiving waters.


14

1.4 Nitrification in saline industrial wastewater 1.4.1 Background Industry is an essential engine of economic growth worldwide and requires adequate resources of good quality water as a key raw material. Global annual water use by industry is expected to rise from an estimated 725 km3 in 1995 to about 1,170 km3 by 2025, by which time industrial water usage will represent 24% of all water abstractions. Industrial use of water increases with country income, ranging from 10% for low-and middle-income countries to 59% for high-income countries (World Bank, 2001). A number of industrial categories (petroleum refining, coke processing, dairy, chemical production, tannery, fish processing) contain significant amounts of nitrogen in their wastewater (Wiesmann 1994). Before discharge to the water body (according to the stricter regulations), almost full nitrogen removal is necessary. This has led to increasing activities in the field of development and optimisation of biological nitrogen removal. Biological nitrogen removal is conventionally achieved by making use of processes of the natural nitrogen cycle, namely through nitrification in an aerobic environment followed by denitrification in an anoxic environment. Biological nitrification-denitrification is the most common processes for nitrogen removal from wastewater; nitrification is the ratelimiting step in biological nitrogen removal, because nitrifying bacteria are growing slowly and are very sensitive to environmental factors (e.g. temperature, pH, dissolved oxygen concentration, toxic and inhibitory compounds) (Antonious et al 1990; Wagner et al 1996; Wagner and Loy 2002). Studies so far have concentrated mainly on domestic wastewater treatment and the results obtained may not be directly applicable to industrial wastewater due to their specific composition (high temperatures, sub-optimal pH values, presence of toxic compounds or high salinity). Thus, special attention for and concern with the design and operation of nitrogen removal for industrial wastewater treatment systems is necessary. The effect of salts on nitrogen removal is a major concern, especially in industrial wastewater treatment. Industries such as pickling, cheese manufacturing, seafood processing, tanning, productions of chemicals and Pharmaceuticals, oil and gas recovery, produce high inorganic salt concentrations in their wastewater. Other sources of saline wastewater include infiltration of subsurface water in the coastal areas into the sewer system, landfill leachates and contaminated ground water and ballast water for marine vessels or offshore installations. In future, waste minimisation practices are expected to generate brines via effective water reuse and recycling schemes. Also the use of saline water for toilet flushing due to the scarcity of fresh water will increase the wastewater salinity that reaches the treatment plants (Campos et al 2002; Dahl et al 1997; Woolard and Irvine 1995; Yu et al 2002). Nitrification is the bottleneck of the nitrogen removal process under salt stress, while denitrification has proved to be more stable under salt stress (Vredenbregt et al 1997; Dahl et al 1997). Studies on the effect of salt on nitrification are contradictory and difficult to be interpreted.

Introduction

15

1.4.2 Nitrification The nitrification process—an important process in the nitrogen cycle in nature—is defined as the biological transformation of reduced forms of nitrogen into nitrite and subsequently to nitrate. Generally, two absolutely different types of nitrification must be distinguished (Schmidt et al 2003): • Lithotrophic nitrification, in which the oxidation of inorganic, reduced nitrogen compounds serves as energy source for growth. Lithotrophic nitrification is carried out by two groups of bacteria, the ammonium-oxidizers and nitrite-oxidizers • Heterotrophic nitrification, in which nitrification is a co-oxidation and does not serve as an energy source. It is carried out by diverse groups of microorganisms (bacteria, fungi, and algae). In natural environments, the chemolithotrophic nitrifiers are the only group of microorganisms producing considerably high amounts of nitrite and nitrate from ammonia. The heterotrophic nitrifiers’ specific activity is estimated to be around 103– 104 times lower than that of lithotrophic nitrifiers and therefore heterotrophic nitrification is of minor ecological significance (Kuenen and Robertson 1994; Richardson et al 1998). Originally, the lithoautotrophic nitrifying bacteria altogether were grouped within one family, named Nitrobacteraceae and composed of two physiologically distinct groups of bacteria that are not phylogenetically related (ammonia-oxidizing bacteria and nitriteoxidizing bacteria). However, phylogenetic investigations made evident that a lot of distinct groups of organisms exist, which are not closely related to each other (Koops and Röser 2001). Cells of both groups are able to aggregate in clusters (flocs), which is common in wastewater treatment plants (Stalely et al 1989). 1.4.2.1 Ammonia oxidisers After the first reports on successful isolation of chemolithoautotrophic ammoniaoxidizing bacteria at the end of the 19th century, researches have continued to investigate their diversity in natural and engineered systems by applying enrichment and isolation techniques. These efforts resulted in the description of numerous species of ammonia oxidisers, now with the modern molecular biological techniques more species have been discovered. Chemolithoaotutrophic ammonia-oxidizing bacteria comprise two monophyletic lineages within the class Proteobacteria (Table 1.6). One group is located within the γ subclass, which contains only the Nitrosococcus oceanus and Nitrosococcus halophilus. The second group belongs to the β subclass, which contains two clusters, the Nitrosospira cluster and the Nitrosomonas cluster (Koops and Röser 2001; Purkhold et al 2000).


16

Table 1.6 The cultured ammonia oxidising bacteria and information on ecophysiological parameters and preferred habitats (Koops and Röser 2001). Species

Ecophysiological parameters

Preferred habitat

Salt requirements

Substrate (NH3) affinity

Nitrosomonas europea Nitrosomonas eutropha Nitrosomonas halophila Nitrosomonas mobilis

halotolerant or moderately halophilic

30–61 µM

sewage treatment plants, eutrophic freshwater and brackish water

Nitrosomonas communis Nitrosomonas sp.I Nitrosomonas sp.II

no salt requirement

14–43 µM

soils (not acid)

Nitrosomonas nitrosa

no salt requirement

19–46 µM

eutrophic freshwater

no salt requirement

1.9–4.2 µM

oligotrophic freshwater natural solis

Nitrosomonas marina Nitrosomonas sp.III Nitrosomonas aestuarii

obligatory halophilic

50–52 µM

marine environment

Nitrosomonas cryotolerans


42–59 µM

βProteobacteria Nitrosomonas ureae Nitrosomonas oligotropha

Nitrosolobus multiformis Nitrosovibrio tenuis Nitrosospira sp.I Nitrosococcus γoceani Proteobacteria Nitrosococcus halophilus

no salt requirement

soils (not acid) Soils, rocks and freshwater


marine environment

Introduction

17

Although the basic metabolism is more or less uniform within the physiologically defined groups of lithoautotrophic ammonia oxidizing bacteria, ecophysiological differences exist between the distinct representatives. Different members of these genera have been found to dominate different wastewater treatment plants or natural ecosystems, but general relationships between the ecological niche and evolutionary position are often still obscure (Schmidt et al 2003). Salt requirement is an ecophysiologically relevant discrimination factor. All isolates of the two species of Nitrosococcus (γ subclass of the Proteobacteria), Nitrosococcus oceani and Nitrosococcus halophilus, are obligately halophilic. The group located in the β subclass of the Proteobacteria, comprises obligately halophilic species and moderately halophilic or halotolerant species, respectively, together with species missing salt requirement or being salt sensitive. Within the genus Nitrosomonas, these differences are well reflected by the pronounced formation of phylogenetic lineages (Koops and Röser 2001). Physiology The physiology of conventional, ‘aerobic’ ammonia oxidizers is not completely understood. Only recently, it was discovered that these organisms also have an anaerobic metabolism. The proteobacterial ammonia oxidizers can obtain their energy for growth from both aerobic or anaerobic ammonia oxidation. Most likely ammonia (NH3) and not ammonium (NH4+) is the substrate for the oxidation process (Suzuki et al 1974; Bock et al 1991). The main products are nitrite under oxic conditions (DO> 0.8 mg O2/L), while under anoxic conditions (DO18 gCl−/L Dincer and SRTmin is 25 Kargi (1999) days, at 0gCl−/L 12 days

Introduction

10–20

NaCl

31–55

NaCl

20–43

SW

EN

no

LA

3,6,12,18 nm*5

27–33

SW

DA

no

LA

NaCl

3,3,12,18

nm

27–33

SW

SA

to 5 g Cl−/L

LA

50%

Nil

SBR effluent

As SBR influent medium(3)

0.035

7.5

7.0

18 gCl−/L 5 SRTmin is 25 days, at 0gCl−/L 12 days

10–20

NaCl

18,30

8

25

SW

EN

no

LA

nm

nm

NO2− accum ulation above 12 gCl−/L

6

31–55

NaCl

3,6,12,18

SW

DA

no

LA

na*5

+

MLSS decreased with increased NaCl

8,9

20–43

NaCl

3,6,12,18

nm

27–33

SW

SA

to 5g Cl−/L

LA

na

+

95%). However, a significant difference among treatment plants in activity of ammonia and nitrite oxidisers was observed.

Table 6.1 Operating data and measured influent and effluent parameters of the 4 full-scale WWTP. Plant Name Parameter

Unit

Type

System Volume Flow Temperature

Hoek van Holland

Heiploeg

Ecco

Seafarm

Domestic

fish processing Industry

tannery Industry

marine aquaculture Industry

AS

SBR

AS

BF

3

(m )

5700

2800

8000

3

21000

440

843

(°C)

14

14–18

22

(m day−1)

100

Nitrification activities in full-scale treatment

127

COD

(mg/L)

420

2050

2175

301.0

Influent TKN

(mg/L)

40

557

519

20.0

7

7.6

7.6

7.3

pH COD

(mg/L)

40

74

117

Effluent TKN

(mg/L)

1.8

12

24

7

7.5

7.6

7.3

pH Reactor

MLSS

(mg/L)

4445

4700

8300

9960

MLVSS

(mg/L)

3200

3390

5870

3540

(ms/cm)

1

10

24

44

(mg/L)

130

3000

8000

16000

(mg/L)

47

68

4000

Conductivity Cl

− 2−

SO4

6.3.1 Activity of ammonia and nitrite oxidisers The results demonstrate a decline in the activity of ammonia and nitrite oxidisers with an increase in salt content (Figure 6.1). The domestic WWTP with the lowest salt level (fresh water, 130 mgCl−/L) had the highest specific activities of ammonia and nitrite oxidisers (4.3 and 2.4 mg-N/gVSS.h, respectively). The lowest specific activities of ammonia and nitrite oxidisers (1.1 and 0.5 mg-N/gVSS.h) were measured at the highest NaCl concentration in the sludge from Seafarm. Under all experimental conditions the specific activity of ammonia oxidisers was always two times higher than that of the nitrite oxidisers.

Figure 6.1 Activity of ammonia and nitrite oxidisers measured at different salt concentrations for sludge collected from 4 full-scale WWTP. The


128

activities were measured under standard conditions: pH 7.5 and T 20°C. 6.3.2 Population diversity of ammonia and nitrite oxidisers Four species of ammonia oxidisers (Nitrosomonas oligotropha, Nitrosomonas europaea, Nitrosococcus mobilis and Nitrosospira sp.) were detected within the collected samples of the 3 industial WWTPs (Table 6.2). No nitrite oxidisers were detected with the two available oligonucleotide probes (NIT3 and Ntspa662).

Table 6.2 Population structure of nitrifying bacteria in industrial full scale WWTPs operated at different salt levels. WWTP

Heiploeg

Ecco

Seafarm

Salt conc. (gCl /L)

3

8

16

Conductivity (ms/cm)

10

24

44

+++

−

+++

Nitrosomonas oligotropha

−

+++

+

Nitrosococcus mobilis

+

−

−

Nitrosospira sp.

+

+

−

Nitrospira sp.

−

−

−

Nitrobacter sp.

−

−

−

−

Ammonia oxidisers: Nitrosomonas europaea

Nitrite oxidisers:

6.4 Discussion 6.4.1 Activity of ammonia and nitrite oxidisers The results showed the adverse effect of salt on the nitrification process in full scale WWTPs. However, the results demonstrated the complexity of comparing nitrification activity of different types of sludge developed under different operational conditions. The operational conditions of industrial WWTPs vary widely and reports on the nitrification performance of their wastwater are very scarce (Wiesmann 1994). Moreover, only few studies distinguish between ammonia and nitrite oxidisers (Hunik et al 1992, 1993; Moussa et al 2003 a, b, c). The decrease in specific activity of ammonia oxidisers with increase of the salt content is in line with the reported results (Dahl et al 1997;


129

Vredenbregt et al 1997; Panswad and Anan 1999 a, b). However, there is much deviation under domestic wastewater conditions (Pollice et al 2002, Salem et al 2003) (Figure.2).

Figure 6.2 Comparison of specific ammonia oxidisers activities (mgN/gVSS.h) measured in the 4 full scale WWTP (taken from figure 1) with the results reported in literature. The results were all calculated to temperature 20 °C, with the formula R20=RT. exp(0.094(T−20)). The specific activity of pure or enriched cultures of nitrifiers under salt-free conditions is 10–15 higher as compared to full scale WWTPs (Table 6.3). This could be explained by the difference in active fraction of ammonia and nitrite oxidisers present in the investigated sludge. The sludge is comprised of inert particulates, biomass of heterotrophs (produced through COD conversion) and nitrifiers (produced through ammonia oxidation), grazers (protozoa and metazoa) and inert particulates produced through the decay of heterotrophs, nitrifiers and grazers. The ratio of each fraction to the total biomass is a function of the operating conditions, such as, solid retention time (SRT), hydraulic retention time (HRT), COD removal and TKN removal.


130

Table 6.3 Reported values of ammonia (AOB) and nitrite (NOB) oxidisers activities of different type of biomass developed in different system and different operational conditions (Temperature, salt level, COD/N ratio, SRT). Specific rate AOB

System*1 Biomass*2 CODin NH4in SRT Reference

Temp.

Salt

°C

gCl−/L

NOB

mg mg NH4– NO2– N/gvss.h N/gvss.h

mg/L

50

days

3.63

22

0.1 FA

DA

22.9

32

0.1 PA

DA

10

32

0.1

5

4.8

32

0.1

14

3.2

32

0.1

24

4.76

28

0.1 LA

3.29

28

2.48

8 Salem et al (2003) 3 Pollice et al (2002)

500

25

10

5

500

25

10

28

10

500

25

10 Panswad and Anan (1999)

2.43

28

20

500

25

10

2.14

28

30

500

25

10

3.5

30

20

SA

42.8

200

2.35

30

20

SA

395.9

35

3

25

10 PF

840

120

2

30

20

840

120

220

20

660

FB

DA

280

mgN/L

PCA

Vredenbregt et al (1997) Dahl et al (1997) Copp and Murphy (1995)

20

PCN

160

20

MC

444

30

LC

PCA

0.3 Hunik et al (1992)

30

LC

PCN

0.3 Hunik et al

1412


131

(1993) 96

58.5

30

0.1 LSBR

EC

10

250

30 Moussa et al (2003c)

58.5

55.7

30

10

EC

10

250

30

17.7

19.6

30

20

EC

10

250

30

14.7

12.2

30

30

EC

10

250

30

3

1.9

30

40

EC

10

250

30

*1

FA=Full-scale Activated sludge WWTP; PA=Pilot-scale Activated sludge unit; LA=Lab-scale Activated sludge unit; PF=pilot-scale Fluid-bed; FB=lab-scale Fed-Batch mode; LC= sterile Lab-scale Chemostat; LSBR=Lab-scale Sequencing batch reactor; PA=Pilot-scale Activated sludge unit *2 DA=Domestic Activated sludge performing nitrification; SA=Salt Adapted activated sludge performing nitrification; PCA=Pure Culture of Ammonia oxidisers Nitrosomona europea; PCN =Pure Culture of Nitrite oxidisers Nitrobacter agilis; MS=Marine Sediment; MC=Nitrosomonas and Nitrobacter in Mixed Culture; EC=Enriched Culture of nitrifying bacteria

The activated sludge model (ASM1) was used to translate the routine operating data into parameters to enable the calculation of the fraction of nitrifiers and consequently to recalculate the actual specific activity of ammonia and nitrite oxidisers (equation 6.1, 6.1). (6.1) (6.2) The ASM1 model was applied to recalculate the actual activity of ammonia oxidisers both in this study and in other reported studies. Ammonia oxidisers: The assessment of the fraction of ammonia oxidisers demonstrates the influence of the above-mentioned operating parameters on the portion of the nitrifiers to the total biomass. This is in agreement with the results reported by Rittmann et al (1999) and Wiesmann (1994). Expressing the activity of ammonia oxidisers in terms of specific activity eliminates the deviation between the specific activity from different systems and makes the results from pure cultures, enriched cultures, pilot scale and full scale WWTPs comparable (Figure 6.3). Moreover, the results of activities of pure and enriched cultures can now be extrapolated to full scale WWTPs. The activity seems to be dependent on the salt concentration, irrespective of the dominant species of ammonium oxidisers, which was also demonstrated by Moussa et al (2003c).


132

Figure 6.3 Specific activity of ammonia oxidisers based on active fraction of the sludge (mg-N/gVSS ammonia oxidisers.h) as a function of salt concentrations. The results of this study are compared with reported data, all recalculated for 20 °C. Nitrite oxidisers: Similar recalculation of the activity of nitrite oxidisers in terms of specific activity agrees with results from pure cultures (Copp and Murphy 1995; Hunik et al 1993) and enriched cultures (Moussa et al 2003c). 6.4.2 Population of nitrifiers Ammonia oxidisers: Both Nitrosomonas oligotropha and Nitrosomonas europaea are normally present together under low salt levels, not only in full scale WWTPs (Daims et al 2001b; Gieseke et al 2001; Juretschko et al 1998; Liebig et al 2001; Purkhold et al 2000), but also in laboratory-scale systems (Nogueira et al 2002; Moussa et al 2003c). At elevated salt levels in the industrial WWTPs only one of the Nitrosomonas species, N. europaea, became dominant, at 3 and 16 gCl−/L (Heiploeg and Seafarm). This is in line with Hovanec and De Longe (1996) and Catalan-Sakairi et al (1996, 1997). Also in enriched cultures with several ammonia oxidising species present at low salt concentrations, N. europaea became dominant when salt was increased above 10 gCl−/L (Moussa et al 2003c). Surprisingly, N. oligotropha was dominant at 8 gCl−/L (Ecco), which could be attributed to the high sulphate concentration (4g SO4−2/L) in this WWTP. It is not surprising that N.europaea becomes dominant under elevated salt levels, because it is halotolerant or moderately halophilic (Koops et al 2001; Wagner et al 1995).


133

Nitrite oxidisers: Despite the fact that nitrite oxidisers were active (oxidation of nitrite to nitrate), neither of the two probes (NIT3 and Ntspa662) were able to detect Nitrobacter sp. or Nitrospira sp. The reason could be that the fraction of these nitrite oxidisers was below the detection limit, since these WWTPs are operated at high sludge age and/or high organic load. Alternatively, nitrite oxidisers might be present, but are not detectable with the two available probes. A number of studies have demonstrated a relationship between the presence of distinct nitrifiers and specific environmental conditions (Kowalchuk and Stephen 2001; McCagig et al 1999; Rittmann et al 1999; Stephen et al 1999). However, a relation between salt and population of nitrifiers cannot be drawn, since within the limited number of WWTPs with elevated salt levels, two different ammonia-oxidising species became dominant. Nevertheless, the use of the model to quantify the specific activity and to correlate this to the presence of specific populations within different systems was promising and could be applied in future to confirm this relation. 6.5 Conclusions • The results showed the adverse effect of salt on nitrification in full scale WWTPs. The domestic WWTP with the lowest salt level (0.13 g Cl−/L) had the highest specific activity of ammonia and nitrite oxidisers (4.3 and 2.4 mg-N/gVSS.h, respectively). The lowest specific activities of ammonia and nitrite oxidisers (1.1 and 0.5 mg-N/g VSS.h) were measured in the WWTP with the highest NaCl concentration (16 g Cl−/L). • It is rather complex to compare between the specific nitrification activity of different types of sludges developed under different operational conditions. This is due to the variation in active fraction of ammonia and nitrite oxidisers present in the investigated types of sludge. • The use of the activated sludge model to calculate the active fraction of nitrifiers and consequently recalculate the actual specific activity of ammonia and nitrite oxidisers makes the results from pure cultures, enriched cultures, pilot scale and full scale WWTPs comparable. • Model application to quantify the actual specific activity validates the previous results of salt on nitrification obtained at laboratory-scale. The ammonia and nitrite oxidisers activity seems to be dependent on the salt concentration, irrespective of the dominant species of ammonium oxidisers. • Both Nitrosomonas oligotropha and Nitrosomonas europaea are normally present under low salt levels, not only in full scale WWTPs but also in laboratory-scale systems. At elevated salt levels only Nitrosomonas europaea became dominant. • A clear identification of the nitrite oxidisers present in the investigated WWTPs was not possible. The reason could be that either the fraction of these nitrite oxidisers was below the detection limit or they were present, but are not detectable with the two available probes.


134

References Antoniou P, Hamilton J, Koopman B, Jain R, Holloway B, Lyberatos G, Svoronos SA (1990) Effect of temperature and pH on the effective maximum specific growth rate of nitrifying bacteria. Water Res. 24:97–101. APHA (1998) Standard methods for the examination of water and wastewater, 20th ed. American Public Health Association/American Water Works Association/Water Environment Federation, Washington D.C. Bock E, Koops HP, Harms H, Ahlers B (1991) The biochemistry of nitryfying organisms. In: Shively JM, Barton LL Variation in autotrophic life. Academic Press, London. Catalan-Sakairi MAB, Wang PC, Matsumura M. (1997) Nitrification performance of marine nitrifiers immobilized in polyester and macro-porous cellulose carriers. Fermentation and Bioeng. 84:563–571. Catalan-Sakairi MAB, Yasuda K, Matsumura M. (1996) Nitrogen removal in seawater using nitrifying and denitrifying bacteria immobilized in porous cellulose carrier. Water Sci. Technol. 34:267–274. Copp JB, Murphy KL (1995) Estimation of the active nitrifying biomass in activated sludge. Water Res. 29:1855–1862. Dabert P, Delgenes J-P, Moletta R, Godon J-J (2002) Contribution of molecular microbiology to the study in water pollution removal of microbial community dynamic. Re/View in Environmental Science and bio/Technology 1:39–49. Dahl C, Sund C, Kristensen GH, Vredenbregt L (1997) Combined biological nitrification and denitrification of high-salinity wastewater. Water Sci. Technol. 36:345–52. Daims H, Brühl A, Amann R, Schleifer K-H, Wagner M (1999) The domain-specific probe EUB338 is insufficient for the detection of all bacteria: development and evaluation of a more comprehensive probe set. Syst. Appl. Microbiol. 22:434–44. Daims H, Purkhold U, Bjerrum L, Arnold E, Wilderer PA, Wagner M (2001b) Nitrification in sequencing biofilm batch reactors: lessons from molecular approaches. Water Sci. Technol. 43:9–18. Gieseke A, Purkhold U, Wagner M, Amann R, Schramm A (2001) Community structure and activity dynamics of nitrifying bacteria in a phosphate-removing biofilm. Appl. Environ. Microbiol. 67:1351–1362. Hellinga C, Schellen AAJC, Mulder JW, van Loosdrecht MCM, Heijnen JJ (1998) The SHARON process: an innovative method for nitrogen removal from ammonia-rich wastewater. Water. Sci. Technol. 37:135–142. Hovanec TA, De Longe E (1996) Comparative analysis of Nitrifying Bacteria associated with freshwater and marine aquaria. Appl. Environ. Microbiol. 62:2888–2896. Hunik JH, Meijer HJG, Tramper J (1992) Kinetics of Nitrosomonas europaea at extreme substrate, product and salt concentrations. Appl. Microbiol. Biotechnol. 37:802–807. Hunik JH, Meijer HJG, Tramper J (1993) Kinetics of Nitrobacter agilis at extreme substrate, product and salt concentrations. Appl. Microbiol. Biotechnol. 40:442–8. Juretschko S, Timmermann G, Schmid M, Schleifer K-H, Pommerening-Röser A, Koops H-P, Wagner M (1998) Combined molecular and concentional analyses of nitrifying bacteriumin activated sludge: Nitrosococcus mobilis and Nitrospira-like bacteria as dominant populations. Appl. Environ. Microbiol. 64:3042–51. Koops H-P, Pommerening-Röser A (2001) Distribution and ecophysiology of the nitrifying bacteria emphasizing cultured species. FEMS Microbiol. Ecology 37:1–9. Kowalchuk GA, Stephen JR (2001) Ammonia-oxidising bacteria: a model for molecular microbial ecology. Annu. Rev. Microbiol. 55:485–529.


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Liebig T, Wagner M, Bjerrum L, Denecke M (2001) Nitrification performance and nitrifier community composition of a chemostat and a membrane-assisted bioreactor for the nitrification of sludge reject waters. Bioprocess Biosyst. Eng. 24:203–210. Manz W, Amann R, Ludwig W, Wagner M, Schleifer KH (1992) Phylogenetic oligonucleotide probes for the major subclasses of Proteobacteria: problems and solutions. Syst. Appl. Microbiol. 15:593–600. Manz W, Amann R, Ludwig W, Vancanneyt M, Schleifer KH (1996) Application of a suite of 16S rRNA specific oligonucleotide probes designed to investigate bacteria of the phylum Cytophaga-Flavobacter-Bacteroides in the natural environment. Microbiol. 142:1097–1106. McCaig AE, Phillips CJ, Stephen JR, Kowalchuk GA, Harvey M (1999) Nitrogen cycling and community structure of β-subgroup ammonia oxidising bacteria within polluted, marine fishfarm sediments. Appl. Environ. Microbiol. 65:213–20. Moussa MS, Lubberding HJ, Hooijmans CM, van Loosdrecht MCM, Gijzen HJ (2003a) Improved method for determination of ammonia and nitrite oxidation activities in mixed bacterial cultures. Appl. Microbiol. Biotechnol. 63:217–221. Moussa MS, Lubberding HJ, Hooijmans CM, Gijzen HJ, van Loosdrecht MCM (2003b) Short term effects of various salts on ammonia and nitrite oxidisers in Mixed Bacterial Cultures. Appl. Microbiol. Biotechnol. (submitted). Moussa MS, Lubberding HJ, Hooijmans CM, Gijzen HJ, van Loosdrecht MCM (2003c) Long Term Effects of Salt on Activity, Population Structure and Floc Characteristics in Mixed Bacterial Cultures of nitrifiers. Appl. Environ. Microbiol. (submitted). Nogueira R, Melo LF, Purkhold U, Wuertz S, Wagner M (2002) Nitrifying and heterotrophic population dynamics in biofilm reactors: effects of hydraulic retention time and the presence of organic carbon. Water Res. 36:469–481. Panswad T, Anan C (1999a) Impact of high chloride wastewater on an anaerobic/anoxic/aerobic process with and without inoculation of chloride acclimated seeds. Water Res. 33:1165–1172. Panswad T, Anan C (1999b) Specific oxygen, ammonia and nitrate uptake rates of a biological nutrient removal process treating elevated salinity wastewater. Bioresource Technol. 70:237– 243. Pollice A, Tandoi V, Lestingi C (2002) Influence of aeration and sludge retention time on ammonia oxidation to nitrite and nitrate. Water Res. 36:2541–2546. Prosser JI (1986) Nitrification, special publication of the society for general microbiology, Oxford IRL Press, Volume 20. Purkhold U, Pommerening-Röser A, Juretschko S, Schmid MC, Koops H, Wagner M (2000) Phylogeny of all recognized species of ammonia-oxidizers based on comparative 16S rRNA and amoA sequence analysis: Implications for molecular diversity survey. Appl. Environ. Microbiol. 66:5368–5382. Rittmann BE, Laspidou C.S, Flax J, Stahl, DA, Urbain V, Harduin, H, van der Waarde JJ, Geurkink B, Henssen MJC, Brouwer H, Klapwijk A, Wetterauw (1999) Molecular and modeling analyses of the structure and function of nitrifying activated sludge. Water Sci. Technol. 39:51–59. Salem S, Berends DHJG, van der Roest HF, van der Kuijl RJ, van Loosdrecht MCM (2003) Fullscale application of the BABE process. Water Sci. Technol. (in press). Sharma B, Ahlert RC (1977) Nitrification and nitrogen removal. Water Res. 11:897–925. Stephen JR, Chang Y-J, Macnaughton SJ, Kowalchuk GA, Leung KT (1999) Effect of toxic metals in indigenous soil β-subgroup proteobacterium ammonia oxidizer community structure and protection against toxicity by inoculated metal-resistant bacteria. Appl. Environ. Microbiol. 65:65–101. Vredenbregt LHJ, Nielsen K, Potma AA, Kristensen GH, Sund C (1997) Fluid bed biological nitrification and denitrification in high salinity wastewater. Water Sci. Technol. 36:93–100. Wagner M, Rath G, Amann R, Koops H-P, Schleifer KH (1995) In situ identification of ammoniaoxidizing bacteria. Syst. Appl. Microbiol. 18:251–64.


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Wagner M, Nogueira DR, Juretschko S, Rath G, Koops H-P, Schleifer KH (1998) Combining fluorescent in situ hybridisation (FISH) with cultivation and mathematical modelling to study population structure and function of ammonia oxidising bacteria in activated sludge. Water Sci. Technol. 31:441–449. Wagner M, Loy A (2002) Bacterial community composition and function in sewage treatment system. Environ. Biotechnol. 13:218–227. Wiesmann U (1994) Biological nitrogen removal from wastewater. Advances in biochemical Engineering/biotechnology 51:113–154.

Chapter 7 Model-based evaluation of the upgrading of a full-scale industrial wastewater treatment plant Previously published as: Moussa MS, Rojas AR, Hooijmans CM, Gijzen HJ, van Loosdrecht MCM (2004) Model-based evaluation of nitrogen removal in a tannery wastewater treatment plant. Accepted for the IWA Conference on wastewater treatment for nutrient removal and reuse. Bangkok, Thailand (January 26–29, 2004).

Abstract Computer modelling has been used in the last 15 years as a powerful tool for understanding the behaviour of activated sludge wastewater treatment systems. However, computer models are mainly applied for domestic wastewater treatment plants (WWTP). Application of these types of models to industrial wastewater treatment plants requires a different model structure and an accurate estimation of the kinetics and stoichiometry of the model parameters, which may be different from the ones used for domestic wastewater. Most of these parameters are strongly dependent on the wastewater composition. In this study a modified version of the activated sludge model No. 1 (ASM 1) was used to describe a tannery WWTP. Several biological tests and complementary physicalchemical analyses were performed to characterise the wastewater and sludge composition in the context of activated sludge modelling. The proposed model was calibrated under steady-state conditions and validated under dynamic flow conditions. The model was successfully used to obtain insight in the existing plant performance, possible extension and options for process optimisation. The model illustrated the potential capacity of the plant to achieve full denitrification and to handle a higher hydraulic load. Moreover, the use of a mathematical model as an effective tool in decision-making was demonstrated.

7 Model-based evaluation of the upgrading of a full-scale industrial wastewater treatment plant 7.1 Introduction The activated sludge system is currently the most widely used biological wastewater treatment process, treating both domestic and industrial wastewater. The process requires a high degree of operational control and management. In order to obtain maximum removal efficiency from the activated sludge plant, the operator must have a full understanding of this complex process. Much research has been done over the last 15 years to understand the behaviour of activated sludge systems using computer modelling. A common language for all modellers in this field regarding concepts and nomenclature is provided by the ASM models developed by the IWA task group (Henze et al 2000). The ASM models have proved to be a useful tool for the dynamic simulation of activated sludge systems treating domestic wastewater. However, application of these models to industrial wastewater treatment plants remains limited. To apply these models to industrial wastewater treatment plants it could be necessary to extend the ASM models by additional kinetic reactions (Nowak et al 1995). The use of models especially in the field of industrial wastewater will support plant operators. The physical, chemical and biological properties of industrial wastewater and their variation in flow and composition make the operation more complicated. Moreover, the model could be used as a quantitative way to predict the effect of different production scenarios on their wastewater treatment plant (van Zuylen 1993), supporting the operator in decision making. The tannery industry is one of the industries generating high amounts of polluted water while the industry has a low profit margin. Therefore, purification of the generated wastewater has a high impact on the overall production costs. The work presented here emphasises on modelling of a tannery activated sludge wastewater treatment plant and its practical application. The main objectives of this study are: • To modify activated sludge model ASM1 to satisfactorily describe the COD and N removal in the tannery wastewater treatment plant; • To evaluate the plant performance using the modified model; • To investigate the required modifications for the plant optimisation and future extensions on plant capacity.

Model-based evaluation of the upgrading

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7.2 Materials and methods 7.2.1 Plant and process description The study covers the wastewater treatment plant of Ecco Tannery Holland B.V., a tannery factory located at Dongen, The Netherlands. The WWTP is in operation since 1987, treating wastewater generated from different steps of the tannery plant and designed for COD and N removal the general lay-out of the WWTP plant is presented in figure 7.1. The configuration of the plant consists of primary and secondary wastewater treatment and sludge treatment. The primary treatment treats the segregated stream containing Crtotal, which is removed by chemical precipitation. The supernatant liquid is pumped to the next step, where it is mixed with the rest of the generated wastewater in a covered equalisation tank (1750 m3). The equalisation tank buffers the dynamic flow generated during the week (5 working days/week) and provides the plant with a minimum flow during the weekend. In the equalisation tank a dose of Fe(OH)3 (iron sludge from a drinking water treatment plant) is mixed with wastewater to remove the S2− compounds. The top gas layer of the equalisation tank is pumped off, washed and used for aeration in the second stage. Eight primary clarifiers (8×50 m3) are used for particulate solid separation at the end of the primary stage.

Figure 7.1 Plant lay-out of the wastewater treatment plant Ecco Tannery B.V. The secondary treatment


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is a conventional plug flow activated sludge system with predenitrification The secondary treatment consists of a conventional plug flow activated sludge system, which is the main focus of this study. The system involves a plug flow reactor of a total volume of 8000 m3. The first part of the reactor (1000 m3) is non-aerated to ensure denitrification while the rest of the reactor is aerated. An internal recycle flow (Qint) supplies the denitrification zone with nitrate. To avoid phosphorus limitation for microbial growth, a flow of 25L/day (Qin3) of H3PO4 of 75% concentration is dosed in the aerated zone. The outlet of the reactor is connected to a secondary settler of 800 m3, where the settled sludge (Qreturn) is pumped to the denitrification zone and the excess sludge (Qex) is pumped to the equalisation tank of the primary treatment. Finally, the treated effluent is pumped via a 37 km force main to the water authority gravity line, which conveys the wastewater to the domestic WWTP Rilland Bath. The sludge treatment deals with primary and chromium sludge. These two different types of sludge are produced in the plant. The primary sludge, which is collected from the primary settlers, is conditioned in a buffering tank (200 m3), de-watered by a filter press, resulting in a sludge cake having a dry content of 30–35%. This sludge cake and the sludge generated from the chromium removal (chromium sludge) are transported by trucks for final disposal. The rejected water resulting from the primary sludge treatment is pumped to the bioreactor (Qin3). The reactors volumes, the hydraulic and operational collected data are summarised in Table 7.1.

Table 7.1 Operational flow data and reactor volumes of WWTP Ecco Tannery B.V. The values printed italic are obtained from mass balances and are used in the model. Flow

Average flow rates m3/d

Reactor

Volume m3

Influent, Qin1

710

Equalisation tank

1750

Influent, Qin2

125 (133)

Primary settler (8×50)

400

Influent, Qin3

0.025

Unaerated zone

1000

Effluent, Qeff

755

Aerated zone

7000

Return Sludge, Return

1920 (650)

Total reactor volume

8000

Internal recycle, Qint

5760

Secondary settler

800

Excess sludge, Qex

80 (90)

Sludge buffering tank

200

7.2.2 Measurements The staff of WWTP Ecco Tannery Holland B.V. provided the routinely collected operational data of the bioreactor and its performance over the year 2000. A detailed sampling and experimental program was conducted in October and November 2000. The pseudo steady state measurements of the WWTP were performed during two sampling


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runs (11 October and 24 November 2000). During each run samples were collected from the influent, return sludge and effluent. The samples were analysed for T (temperature), pH, Alkalinity, DO (dissolved oxygen), CODtot (total COD), CODS (COD of the microfiltrated fraction, 0.45mm pore diameter), NH4–N, NO3–N, TKN (Total Kjeldal Nitrogen), Ptot (total phosphorus), VSS (Volatile Suspended Solids) and TSS (Total Suspended Solids). In addition, six different sampling points over the length of the bioreactor were defined and sampled during the second run (24 November 2000). These six sampling points were used to describe the hydraulic regime and biological conversion as function of the reactor length. The average values of the routinely collected data for the year 2000 and the average measurements of the sampling program are presented in Table 7.2. Several biological batch tests were performed at the UNESCO-IHE laboratory to determine the influent and sludge characteristics (Ekama et al 1986; Orhon et al 1999a, b). The tests were performed at controlled temperature 20°C, pH of 7.5±0.05 and under aerobic and anoxic conditions. Nitrification batch tests were performed under aerobic conditions, in which NaNO2 and NH4 were consequently injected (Moussa et al 2003a). This test allows measuring the kinetic parameters of nitrite and ammonia oxidisers and was used for model calibration.

Table 7.2 Measured influent and effluent and the influent composition required for the model of WWTP Ecco Tannery B.V. The yearly average values are printed in Italic, these values were used to calculate steady-state influent composition. Measurements

Model Influent Composition

Value

Value

Description Influent Effluent

units

Description

Symbol Dynamic Steadystate

units

Total COD, CODTotal

2525 (2920)

167 (190)

gCOD/m3

Soluble compounds

Soluble COD, CODS

1785

143

gCOD/m3

Dissolved oxygen

SO2

0.3

0.3

gCOD/m3

Total N–Kj

488 (515)

7 (11)

gN/m3

Readily biodegradable COD

SS

840

972

gCOD/m3

Soluble N– Kj

454

6

gN/m3

Soluble inert COD

SI

177

205

gCOD/m3

Ammonium, NH4+

438

10

gN/m3

Ammonium

SNH4

438

448

gN/m3

Nitrite, NO2−

0

0 (0.1)

gN/m3

Nitrite

SNO2

0

0

gN/m3

Nitrate NO3−

0

99 (50)

gN/m3

Nitrate

SNO3

0

0

gN/m3


Total phosphorus, PTotal

0.4

3.3

gP/m3

Ortho phosphate, PO42−

0.1

2.8

gP/m3

Total Chromium, Cr3+

0.23

0.2 (0.1)

gCr/m3

Calcium, Ca2+

200

320 (400)

gCa/m3

Particulate compounds

Chloride, Cl−

8066

7960 (7380)

gCL/m3

Sulfate, SO42−

3700

3900 (3100)

Total Suspended Solids, TSS

1593

Volatile Suspended Solids, VSS Alkalinity Temperature pH Dissolved Oxygen, DO

Alkalinity

144

SAlk

50

50

mole HCO3/l

Inert particulate COD

XI

454

525

gCOD/m3

gSO4/m3

Slowly biodegradable COD

XS

1060

1226

gCOD/m3

500

g/m3

Heterotrophics

XH

0

0

gCOD/m3

255

127

g/m3

Ammonia oxidisers

XNH4

0

0

gCOD/m3

2030

570

gCaCO3/m3 Nitrite oxidisers

XNO2

0

0

gCOD/m3

22

22

°C

9 (9.4)

7.5 (7.1)

0.3

67

gO2/m3

7.2.3 Process model (selection and adjustment) The simulations of the secondary treatment of Ecco Tannery Holland B.V. WWTP were performed with AQUASIM® (Reichert 1998), a computer software package used for simulation. The process was modelled according to the flow scheme in Figure 1, with the hydraulic and operational parameters as presented in Table 1. The plug flow condition of the bioreactor was modelled as six completely mixed stirred compartments in series with an internal recycle flow. In the model the secondary clarifier separates solids and water ideally. The amount of suspended solids discharged in the effluent was considered in the sludge age (SRT) calculation. Oxygen concentrations in each compartment were controlled with a PI-controller in accordance to the measured values. A modified version of the activated sludge model No. 1 (ASM 1) proposed by the IWA task group (Henze et al 2000) was used to calculate the biological conversions in


145

each compartment (Appendix 7.1 and 7.2). The main modification incorporated in ASM1 to simulate the WWTP was that the nitrification was considered as a two-step process (Nowak et al 1995). The reason for this is that ammonia oxidisers grow faster than nitrite oxidisers at temperatures above 15–20 °C (Hellinga et al 1999). Moreover, inhibiting compounds present in industrial wastewater might lead to an adverse effect on one or both steps of the nitrification process. Thus describing the nitrification in two steps enables the identification of any inhibition and the detection of partial nitrification (NO2 accumulation). 7.2.4 Influent measurement and characterisation For the use of the model to simulate the WWTP a detailed wastewater characterisation to determine the model components is required. Laboratory tests involving biodegradation were conducted to determine the influent characteristics. The total influent COD can be described as: CODtotal=SS+SI+XS+XI The readily biodegradable COD (SS) was determined by two different approaches using an aerobic and anoxic batch test as described by Ekama et al (1986). The soluble inert COD (SI) was determined according to the approach suggested by Orhon et al (1999a, b). The Influent XI fraction (XI/CODtotal) ratio was estimated as a result of model calibration fitting the solid COD balance as proposed by Meijer et al (2001) and consequently the rest will represent the XS fraction. Average influent measurements and the calculated model influent compositions are presented in Table 7.2. 7.3 Balancing operational data and measurements 7.3.1 Estimation of sludge age, Q recycling and Qin2 For a reliable simulation study the sludge age (SRT) should be known within 95% accuracy (Brdjanovic et al 2000; Meijer et al 2001). Therefore a check on the SRT (or sludge production) is strongly recommended. For the evaluation of sludge production the overall phosphorus balance was used as proposed by Nowak et al (1999). Three balances were formulated (see Figure 7.2): the overall P balance (equation 1), the overall flow balance (equation 2) and the P balance over the settler (equation 3). Qin1Pin1+Qin2Pin2+Qin3Pin3=QeffPeff+QexPex (1) Qin1+Qin2+Qin3=Qeff+Qex (2) (Qin1+Qin2+Qin3)Pr=QeffPeff+(Qreturn+Qex)Pex (3)


146

Figure 7.2 Phosphorous and overflow balance over the bioreactor and settler. The mass balances are presented in equation (1)–(3). Concentration and flow are used from table 1, 2 and 3. indicate the sampling points. The yearly average measurements of Qin1, Qin3 and Qef in addition to the two runs average measurements of Pin1, Pin2, Pin3 (chemical P addition) and Pex were used to evaluate the Qin2, Qreturn and Qex and consequently the SRT. The calculated Qin2, Qex and SRT (70 days) were in agreement with the recorded value of the plant. However, the balanced Qreturn (650 m3/d) was found to be inconsistent with the reported value (1920 m3/d). The reported value is clearly inconsistent with the measured sludge concentrations in the reactor and return sludge, indicating that the reported value is wrong. Therefore we used the balanced value in simulating the treatment plant. 7.4 Model calibration and simulation After the determination of the main operational parameters and the influent characterisation, the model of the WWTP was calibrated. A step-wise approach was applied as proposed by Meijer et al (2001). First the solids were fitted (P, COD and TKN) on the basis of yearly average measurements. Next the nitrification and denitrification were calibrated on the basis of yearly average measurements. 7.4.1 Calibration of the solids The solids balance is a non-conserved balance. An incorrectly assumed COD load or sludge production will generally be compensated by the simulated oxygen consumption of the process. Because the SRT is fixed according to the PTOT balance, the sludge-COD concentration in the process is mainly determined by the influent XI/CODtotal ratio (fXIin). Inert COD (XI) accumulates in the process. Increasing fXIin therefore leads to increasing the COD in the process and vice versa. By adjusting the influent ratio fXIin to 0.18 the model described the measured MLVSS in the reactor (using a conversion factor of 1.4


147

gCOD/gMLVSS). When fXSin is used to fit the COD balance, all model uncertainties related to the production of XI and the influent characterisation of XS and XI are lumped in the influent fXIin fractionation. As a result of adjusting the fXIin fraction, the model predicted accurately the soluble COD in the effluent and total P and TKN in the bioreactor. 7.4.2 Calibrating nitrification and denitrification Nitrification and denitrification were calibrated on the basis of yearly average measurements. Adjusting the oxygen half saturation coefficient of the ammonia oxidisers and heterotrophic biomass was used to fit the nitrification and denitrification in the bioreactor according to Meijer et al (2001). To simulate the measured effluent ammonium and nitrate concentration, a value of 1 mgO2/l and 0.75O2/l and respectively. The calibrated values of and are linked was used for to oxygen diffusion limitation within the sludge floc and oxygen concentration gradients in the tanks caused by non-ideal mixing, processes that are not accounted for in the and are expected to be simulations. Because of differences in mixing intensities, different for each compartment but the same values were used in all compartments for simplicity. Since these values are most sensitive in the anoxic compartment this will not affect the outcome of the simulations significantly. The nitrification batch tests were used in this study as a last step in the calibration procedure. Ammonia and nitrite oxidising activities in the batch test were not predicted very well by the model; observed ammonia conversion and nitrite conversion rates were approximately 20% lower than predicted. Due to the fact that the WWTP is under-loaded, these differences could not be observed when simulating the full-scale yearly average values. There are three possible reasons for this lower experimental nitrification conversion rates in the batch tests than predicted by the model: either the amount of nitrifiers predicted by the model in the plant is too high or the growth rate in the model is too high or both. This means that re-calibration was required. Since the decay rate is the most uncertain parameter we choose to calibrate on this coefficient. Increasing the decay rates by 30 % (from 0.15 and 0.10 day−1 to 0.20 and 0.13 day−1 for the ammonia and nitrite oxidisers respectively) resulted in a good fit of the predicted data to the measurements. When this newly calibrated model was used for the full-scale simulation of the treatment plant, the nitrogen content in the effluent did not change. In a separate test similar results were obtained when the growth rate was reduced because of the high correlation of the parameters decay rate and growth rate. 7.5 Model validation Model validation was performed via validating the capacity to predict the measured concentrations of NH4+, NO2− and NO3− along the bioreactor length using the dynamic influent data. The recorded average daily influent flow during the period 13–26 November 2000 (Figure 7.3) was used as input flow for the model. The simulated value of NH4+, NO2− and NO3− of day 24 November (the 12th day of the 14 days dynamic simulation period) were compared with the values obtained from sampling over the


148

length of the bioreactor conducted on the same day. The model provides a rather accurate prediction of the NH4 and NO3 along the bioreactor (Figure 7.4).

Figure 7.3 Measured influent flow (Qin1) during 13–26 November 2000, on day 11/24/2000 samples were taken and used for model validation (see Figure. 7.4)

Figure 7.4 Measurement results in terms of nitrogen over the length of the bioreactor of WWTP Ecco (markers) on 11/24/2000. The curves are calculated by the calibrated model.


149

7.6 Model application This study was conducted in the first year of changing the ownership of the company. The new owner performed various viability studies on increasing the hide production and application of different tannery process and operational strategies. The main question, which arose, was whether the existing WWTP could cope with these future requirements. Based on the plant performance and effluent quality, an additional volume requirement (25% of the bioreactor volume) was proposed by the operator. The old equalisation tank was selected to be the additional volume, and to be operated as an anoxic reactor. This upgrading aimed to increase the plant capacity and to achieve full denitrification. Evaluating the upgrading concept for the WWTP was the main focus of our modelling study. The validation step demonstrated the capability of the proposed model to describe the WWTP correctly under both steady state and dynamic conditions. Hereafter, the model was used as a tool to obtain insight in the existing plant performance, possible extension and ways of process optimisation. 7.6.1 Evaluation of the existing plant performance and possible extension The model visualised the existing conditions and the potential capacity of each process involved in the purification system. Nitrification as the most sensitive process was accomplished within 60% of the total bioreactor volume under both steady state and dynamic conditions (Figure 4). Moreover, denitrification seemed incomplete (only 80%). The same effluent quality was also predicted even up to a load increase by a factor two. This illustrates that the WWTP Ecco has been under-loaded and that the bioreactor volume is not the limiting factor if the load is expanded. It is possible that the aeration capacity and the sludge treatment units will be limiting factors in the treatment process in possible expansion. Despite the high concentration of NO3 in the effluent (50 mg-N/l) the present set-up of the plant has a high denitrification potential. Within the existing process configurations better effluent quality with less operational costs could be achieved via process optimisations. 7.6.2 Process optimisations One of the major achievements of plant optimisation is enhancing the denitrification activity to achieve full denitrification in the system. The model was used to quantify the main limiting factors, which hinder the denitrification. Anoxic zone volume, internal recirculation flow and the availability of easily degradable COD were investigated as limiting factors. Different modifications were simulated under steady-state conditions to study their effect on the realisation of full denitrification. COD limitation seemed to be the most crucial factor for the denitrification limitation. Different modifications were simulated under steady-state conditions to study their effect on the realisation of full denitrification. Introducing the reject water (Qin2) into the denitrification zone in addition to increasing the denitrification zone (up to 4000 m3) and increasing the internal recycle flow (up to 10000 m3/h) will directly lead to full denitrification. However, increasing the internal recycling flow should be carefully performed to avoid a highly aerated recycling flow to the denitrification zone. Therefore, oxygen should be monitored. Improving the


150

mixing condition in the denitrification zone was recommended to improve the denitrification capacity of the system. This results in a reduction in oxygen consumption, reduction in additional alkalinity to be provided and a reduction in the effluent pollutant (NO3) concentration. This would therefore, result in a reduction in the over-all operational costs. 7.7 Discussion The influent characterisation results of WWTP wastewater do agree with the reported values of the tannery wastewater (Orhon et al 1999a). The modified ASM1 model proposed in this paper for COD and N removal described the performance of the WWTP An accurate and correct well with the adjustment of only two parameters description of the system configurations, balancing the operational data with the measurements to accurately calculate the SRT and the use of the stepwise calibration proposed by Meijer et al (2001) simplified the complexity of the model calibration. The use of batch tests for model re-calibration was useful because the plant was under loaded and therefore the effluent concentrations profiles were not sensitive for the rate parameters. Model validation under dynamic conditions is of great significance in industrial WWTP because the plant is usually working under highly dynamic conditions in comparison with domestic WWTP. The observed reduction in ammonia and nitrite oxidising activity is attributed to the presence of salt (7.5 g Cl−/L). This decline in the activity for ammonia and nitrite oxidisers is in agreement with earlier results of salt inhibition effects on nitrifying sludge (Moussa et al 2003b). The use of a higher decay rate of nitrifiers to mathematically describe the salt impact simplifies the model description of salt inhibition (only one parameter to calibrate). Other parameters for calibration could have been chosen but the decay rate was considered as the most uncertain coefficient in the model. The potential use of the model was illustrated in the evaluation of the upgrading of the WWTP. During this evaluation the bioreactor capacity and process configuration were investigated. The model provided a better and quantitative understanding of the plant operation and treatment process. The investigation of several modifications to reach further plant optimisation was very time-effective and cheap (Salem et al 2002). Because it gives quantifiable results, modelling supported the decisions to be taken with respect to the plant extension. It was originally proposed to increase the plant volume (25% with the use of old equalisation tank) to cope with anticipated load expansion. The simulations clearly indicated that with the present system and future increase in load good effluent quality can be reached even combined with an increased denitrification. An additional application of the model is the usage by the plant manger or the plant staff. These people use the model in a different way, namely, as a tool to quantitatively predict the effect of certain decisions on the treatment process (Salem et al 2002). This application of the model is very useful in case of industrial WWTP, like the tannery studied here. The cost of treating the wastewater generated during the production process has a crucial impact on the overall production costs. This also increases the awareness of the impact of each pollutant term in each part of the process and stimulates the practice of waste minimisation to have an environmentally friendly production.


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7.8 Conclusions Activated sludge models commonly applied for domestic wastewater treatment plant could also be used for industrial WWTPs if the following steps are carefully considered: 1. An accurate description of the system configurations; 2. Balancing the operational data with the measurements to accurately calculate the main important reactor input parameters (flow rates and SRT); 3. Selection of model process and components which are significant and dynamic in this system configuration; 4. Complementary analyses to assess the wastewater and the sludge characterisation; 5. Stepwise calibration under steady-sate conditions and finally model validation under dynamic conditions. The modified ASM1 model proposed in this paper for COD and N removal proved to be able to describe the performance of Ecco Tannery Holland B.V. WWTP wastewater treatment plant. The model was successfully used to evaluate and optimise the plant performance. In addition it was demonstrated that the model could be used by the plant manager to support his decisions quantitatively resulting in saving time and money. References Brdjanovic D, van Loosdrecht MCM, Versteeg P, Hooijmans CM, Alaerts GJ, Heijnen JJ (2000) Modelling COD, N and P removal in a full-scale WWTP Haarlem Waarderpolder. Water Res. 34:846–858. Ekama GA, Dold PL, Marais GvR (1986) Procedures for determining influent COD fractions and the maximum specific growth rate of heterotrophs in Activated sludge systems. Wat. Sci. Tech. 18:91–114. Hellinga C, van Loosdrecht MCM , Heijnen JJ (1999) Model based designed of a novel process for ammonia removal from concentrated flow. Mathematical and Computer Modelling of Dynamic Systems 5:351–371. Henze M, Gujer W, Mino T, van Loosdrecht MCM (2000). Activated sludge models ASM1, ASM2 and ASM3. Scientific and Technical Report, IWA Publishing, London. Meijer SCF, van Loosdrecht MCM, Heijnen JJ (2001) Metabolic modelling of full-scale biological nitrogen and phosphorus removing WWTPs. Water Res. 35:2711–2723. Moussa MS, Lubberding HJ, Hooijmans CM, van Loosdrecht MCM, Gijzen HJ (2003a) Improved method for determination of ammonia and nitrite oxidation activities in mixed bacterial cultures. Appl. Microbiol. Biotechnol. 63:217–221. Moussa MS, Lubberding HJ, Hooijmans CM, van Loosdrecht MCM, Gijzen HJ (2003b). Short term effects of various salts on ammonia and nitrite oxidisers in Mixed Bacterial Cultures. Appl. Microbiol. Biotechnol. (Submitted). Nowak O, Svardal K , Schweighofer P (1995) The dynamic behaviour of nitrifying activated sludge systems influenced by inhibiting wastewater compounds. Wat. Sci. Tech. 31:115–124. Nowak O, Franz A, Svardal K, Muller V, Kuhn V (1999) Parameter estimation for activated sludge models with the help of mass balances. Water Sci. Tech. 39:113–120. Orhon D, Ates E, Ubay Cokgor E (1999a) Modelling of activated sludge for tannery wastewaters. Water Environment Research, 71:50–63.


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Orhon D, Ubay Cokgor E, Sozen S (1999b). Experimental basis for the hydrolysis of slowly biodegradable substrate in different wastewaters. Wat. Sci. Tech. 39:87–95. Reichert P (1998) AQUASIM 2.0 Computer Program for the Identification and Simulations of Aquatic Systems. Swiss Federal institute For Environmental Science and Technology (EWAG), Dübendorf, Switzerland. Salem S, Berends D, van Loosdrecht MCM, Heijnen JJ (2002) Model-based evaluation of a new upgrading concept for N-removal. Wat. Sci. Tech. 45:169–176. van Zuylen HJ (1993) From scientific computation to decision support. Knowledge-based system 6:3–10.

Appendix to Chapter 7

APPENDIX 7.1 Stoichiometry matrix of the activated sludge model (modified ASM1) applied for the industrial WWTP Ecco Tannery B.V.

APPENDIX 7.2 Process kinetics of the activated sludge model (modified ASM1) applied for the industrial WWTP ECCO Tannery B.V.

Chapter 8 Evaluation and Outlook

8 Evaluation and Outlook 8.1 Introduction The aim of this research was to achieve a better understanding of nitrification under saline conditions. The research was carried out in two phases. In the first phase, laboratory scale activities were conducted to obtain fundamental data to determine the relationship between salinity and nitrification. In the second phase the results collected from the laboratory experiments were compared and validated with the results collected from full-scale treatment plants. Modelling was employed in both phases to provide a mathematical description for salt inhibition on nitrification and to facilitate the comparison. The research has lead to an improved understanding of the effect of salinity on nitrification, while subjects for further research were also identified. The research findings and challenges are described in more detail below. 8.2 Detection of nitrification inhibition In industrial activated sludge wastewater treatment there is always the risk of inhibitory compounds in the influent due to spilling or other incidents in the industry. An effective control in order to maximise the use of the plant volume is needed. Therefore an early detection of inhibitory compounds is essential. The nitrification process is one of the most sensitive processes within modern WWTPs and any disturbance might lead to washout of nitrifiers. Moreover, it requires a long time period before these microorganisms are fully re-established in the plant. Thus, a high priority is always given to the nitrification parameters when control strategies are applied to the activated sludge system designed for COD and N removal. One of the achievements of this study was the development of a simple and reliable method to measure the activity of ammonia and nitrite oxidisers separately in mixed bacterial cultures. The main advantage of this method is the possibility to differentiate between the activities of ammonia and nitrite oxidisers without the use of metabolic inhibitors. The method can be applied to measure the effects of specific inhibitors on the activity of both groups of nitrifiers. It can, therefore, also be used as a detection method for early diagnosis of nitrification problems. Incorporating the method in the existing protocol of the automatic respirometer/on-line is promising and could be a helpful tool to optimise nitrogen removal processes.

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8.3 Nitrification under salt stress Initially, short-term effects of several salts on nitrification were studied. Different types of salts appeared to have different inhibition effects on the ammonia and nitrite oxidisers. Therefore, formulas were developed to unify these differences in inhibition effects on both ammonia and nitrite oxidisers. These formulas can be used in the design, operation and control of the WWTPs that operate under salt stress. These findings facilitate the mathematical description of salt inhibition on nitrifiers from a process-engineering point of view, although the real inhibition mechanisms are not yet clearly understood. Investigating the actual inhibition mechanisms in depth (cell membrane, energy requirement, enzyme inhibition) might not be required for optimal process design. Nevertheless, it could be an interesting subject for further research, since it might lead to a more accurate description of inhibition. Acclimatisation is a term, which is commonly used to describe the recovery of bacterial cultures when exposed to a certain inhibitor over extended periods of time. Acclimatised cultures are recommended for seeding a system that will be operated under similar conditions to shorten the start-up period. Salt acclimatisation was investigated within the course of this research. The effect of salinity on the activity, the composition of nitrifying populations and floc characteristics was observed. The main finding was that acclimatised and non-acclimatised nitrifying sludge were behaving similarly with respect to activity and population selection in response to different salt concentrations. This means in practice that sludge from a conventional domestic nitrifying WWTP can be used as a source for nitrifiers to become adapted to salt stress via gradual adaptation. This avoids the necessity of seeding a system that is expected to operate under salt stress with salt acclimatised sludge. The monitoring of the activity and population composition of a full-scale reactor during the start-up period, after seeding according the proposed approach could be an interesting subject for further research. The selection of Nitrosomonas europaea and Nitrobacter sp. as a result of gradual increase in salinity was shown by the re-growth of only these two species after lowering the salt concentration to zero values. This result shows the high resistance of these species to salt stress. Interestingly, the systems were able to fully recover their nitrification activity with reduced diversity in the population of nitrifiers. These findings support the hypothesis that the conversion rate depends on environmental conditions and not on the type of species. In other words, different types of nitrifiers behave kinetically similar under similar environmental conditions. This hypothesis remains to be verified and could be an interesting topic for further research. 8.4 Role of predators in nitrifying activated sludge systems The current design concepts of activated sludge systems consider bacteria as the sole active biomass. The activity of all other microbial community members (protozoa, metazoa, bacteriophages etc.) is hidden in a simple decay process responsible for the reduction of active biomass. This decay process is the sum of several independent


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processes like maintenance, lysis due to bacteriophage infection and predation. Commonly, protozoa in activated sludge systems were used as bio-indicators linked to process performance and effluent quality and crucial in obtaining a good effluent quality. In this research the presence of protozoa and metazoa in the SBRs (Sequencing Batch Reactors) was monitored in order to provide information to understand their role in these systems. A mathematical model describing the interaction between nitrifiers, heterotrophs and predators has been developed. The model successfully describes the performance of SBRs and predicts the fraction of active biomass (ammonia oxidisers, nitrite oxidisers and heterotrophs). The developed model is a first step in generating a better understanding of the role of predators in activated sludge systems and it opens the black box in which all predator processes are put under the decay process. Further investigations are needed to verify the laboratory scale findings under full-scale activated sludge conditions. Moreover, further development of microbiological measurements able to quantify the active biomass for each individual group of organisms is highly needed. These developments can help in bridging the gap between the microbiologist, biochemists, engineers and process modellers. 8.5 Application of activated sludge model in industrial WWTPs Activated sludge models have been successfully applied in the last two decades for domestic wastewater treatment plants. However, their application to industrial WWTPs is still limited for several reasons. The main reason is the complexity of industrial WWTPs in terms of dynamic flow, waste composition, etc. Another important reason relates to the lack of information of model parameters under extreme conditions prevailing in industries (pH, temperature, toxicant presence). Besides, large variations exist between similar industries, which make the treatment process specific for each individual plant. All these factors have limited the applicability of the existing activated sludge model and the transfer of existing experiences between industries. Therefore, the use of large reactor volumes, excess aeration end extra chemical dosages is commonly practised in industries to ensure compliance with effluent quality requirements. Mathematical modelling has shown to increase the confidence of operators in WWTP design and thereby limits the use of an unnecessary high safety factor for new or upgraded treatment plants. This will directly be translated into lower treatment costs. In this study a modified version of the activated sludge model No. 1 (ASM 1) was applied under static and dynamic conditions. Firstly, the steady state model was used to calculate the actual specific activity of ammonia and nitrite oxidisers of different WWTPs investigated in the course of this research. The routine operating data were used to assess the fraction of nitrifiers in the sludge. The results from different types of nitrifying populations (pure cultures, enriched cultures) and different scale reactors (lab scale, pilot scale, full scale WWTP) were compared. The steady state model confirmed the behaviour of nitrifiers under salt stress and validated the results obtained from laboratory scale (this study) to be interpreted on full scale. The use of the model to quantify the specific activity and to correlate this to the presence of specific populations within different systems was promising and could be applied in future research. Secondly, the dynamic

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model was used to describe a full-scale tannery WWTP. The model was used to obtain insight in the plant performance, possible extension and options of upgrading. The model illustrated the potential of the plant to have a better effluent quality and to handle a higher hydraulic load with simple modifications. The model was used as an effective tool to visualise and quantify the plant conditions. The results demonstrate the application of the model as a decision support tool especially in cases when industrial treatment plants are exposed to substantial fluctuations in production. Based on the experience gained in the course of this research it is concluded that more efforts should be made to motivate designers and operators of industrial WWTPs to become familiar with activated sludge models. A good approach would be to develop a standard influent characterisation procedure for industrial wastewater similar to the one successfully developed and applied for domestic wastewater. Moreover, the model parameters need to be extended to satisfy the operational conditions in industries (e.g. high temperature and long SRT). Disturbance factors such as pH, salinity, etc could be incorporated in the model. All this in addition to producing simple and reliable detection sensors could promote the application of activated sludge modelling in the industrial field. In conclusion, the research has lead to an improved understanding of the effect of salinity on nitrification. The results obtained within the course of this research can be used to improve the sustainability of the existing WWTPs operated under salt stress. The findings also form a guideline for more economical and sustainable design and start up of WWTPs dealing with salt in future.

Samenvatting Biologische verwijdering van stikstof via nitrificatie en denitrificatie wordt zeer algemeen toegepast in de afvalwaterzuivering. Nitrificatie, de snelheidsbepalende stap in het hele proces, blijkt echter moeilijk te sturen, omdat nitrificerende bacteriën enerzijds traag groeien en anderzijds erg gevoelig zijn voor allerlei omgevingsfactoren (temperatuur, pH, zuurstofconcentratie, remmende stoffen). Tot nu toe heeft wetenschappelijk onderzoek zich voornamelijk gericht op nitrificatie in huishoudelijk afvalwater en er is op dit gebied dan ook veel kennis en practische ervaring opgedaan met als resultaat de algemene toepassing van biologische stikstofverwijdering uit huishoudelijk afvalwater. Helaas zijn deze positieve ervaringen niet direct toepasbaar op industrieel afvalwater vanwege de specifieke samenstelling ervan. Veel industrieën hebben ook nog te maken met hoge zoutconcentraties in hun afvalwater. Het zoutgehalte van industrieel afvalwater zal in de toekomst verder toenemen, omdat de politiek een efficiënter gebruik van water eist. Hogere zoutconcentraties zullen een negatieve invloed hebben op de nitrificatie. Een beter begrip van de effecten van zout(en) op de nitrificatie is dan ook hard nodig. Weliswaar is bekend dat de activiteit van zowel ammonium- als nitrietoxideerders afneemt bij hogere zoutconcentraties, maar veel details zijn nog niet bekend: Geven alle zouten dezelfde remming, wat is de maximaal getolereerde zoutconcentratie, zijn ammonium- of nitietoxideerders het meest gevoelig, is adaptatie aan zout mogelijk, zijn sommige nitrificeerders beter bestand tegen zoutbelasting dan andere? Het begrijpen van de effecten van zout op de nitrificatie staat centraal in dit proefschrift, waarbij de bovengenoemde vragen de leidraad vormen voor het onderzoek, dat in twee fasen is uitgevoerd. De proeven op laboratoriumschaal (eerste fase) leverden fundamentele gegevens over de invloed van zout op de nitrificatie; in de tweede fase werden de resultaten van de laboratoriumproeven vergeleken met en gevalideerd met proeven in industriële afvalwaterzuiveringen op praktijkschaal. Computer modellen zijn gebruikt in beide fasen, zowel voor een goede mathematische beschrijving van de relatie tussen zout en nitrificatie als om de onderlinge vergelijking te vergemakkelijken. In de eerste fase is er een methode ontwikkeld om simultaan de activiteit van ammonium- en nitrietoxideerders in mengpopulaties van bacteriën te meten; deze methode is gedurende het hele onderzoek toegepast om de remmende effecten van zout op zowel ammonium- als nitrietoxideerders vast te stellen. Ammonium- en nitrietoxideerders reageerden verschillend op de aangeboden zouten. Vervolgens werden de lange-termijn effecten van aanpassing aan 10 g NaCl per liter bestudeerd, niet alleen op de activiteit van de nitrificeerders, maar ook op de samenstelling van nitrificerende bacteriepopulaties en op de eigenschappen van het bacterieslib. Er werd geen verschil gevonden tussen de wel en niet aan 10 g NaCl geadapteerde nitrificeerders; in beide gevallen werd bij 40 g NaCl een remming van 95% gevonden. De enig overgebleven

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nitrificeerders bij hoge zoutconcentraties waren Nitrosomonas europaea en Nitrobacter sp. Hogere zoutconcentraties hadden een betere bezinking van het bacterieslib tot gevolg. Al bij lage zoutconcentraties werden de aanwezige bacterie-etende protozoën en metazoën gedood. Op basis van dit effect is een mathematisch model opgesteld om de interacties tussen nitrificeerders, heterotrofe bacteriën en bacterievore organismen aanschouwelijk te maken. Tijdens fase 2 werd duidelijk dat het effect van zout op de activiteit van nitrificeerders in huishoudelijk en industrieel afvalwaterzuiveringen niet verschilde van de effecten op laboratoriumschaal. Het bestaande actief-slib model (ASM1) is aangepast om CZV en stikstofverwijdering in industriële afvalwaterzuiveringen onder zoutbelasting te simuleren. Het onderzoek heeft geleid tot een beter begrip van het effect van zout op nitrificatie. De verkregen resultaten kunnen worden gebruikt om de duurzaamheid van de huidige afvalwaterzuiveringen met verhoogd zoutgehalte te verbeteren.

Acknowledgments Thanks to Allah the exalted, the most merciful, for giving me the strength and persistence to keep going with this research even during the most difficult moments. May Allah accept this work and count it as a good deed. My deep thanks to my country Egypt where I grew up and had the first lessons and experiences in my life, I hope I can pay it back some day. I would like to express my thanks to my promoters: Prof. Huub Gijzen who stimulated and supported the formulation of my PhD joint project. Huub, you had an important role in guiding and challenging not only during the research but also after the completion; Prof. Mark van Loosdrecht for his ideas in the earlier stage of the research, valuable feedback, enthusiasm and friendship. Mark, you were always inspiring, educating and available when I needed you. I am also grateful to my supervisors: Tineke Hooijmans who supported me especially during the difficult initial period of this study and for her efforts in establishing the project; Henk Lubberding who guided me during the research, inspired my microbiological experience and spent lots of time helping in finalizing this work. I would like to acknowledge the MSc students, Samir Ibrahim, Deepthi Sumanasekera, Alejandro Rivera Rojas, Orleans Garcia, Aboubakar Gomina, Jochem Smit, Yan Song, Said Rehan, Akram Botorous and Hala Elsadig who were involved in this study for their valuable contribution. I would like to express my appreciations for the endless help and support of the laboratory staff at UNESCO-IHE: Fred Kruis, Frank Wiegman, Kees Bik, Peter Heerings,. Special thanks to the staff of Kluyver Laboratory of TU Delft: Sjaak Lispet, Stef van Hateren and Udo van Dongen for their continuous help during this study. Great thanks to my colleagues at IHE Saleh and Saber for their continuous encouragement and inspiring discussions on our related research topics. Special thanks to my friend M.Fiala for all what he did and what he is still doing for me. These acknowledgments would not be complete without expressing my gratefulness to GOSD for expanding my background as a structural engineer with environmental experience and also my appreciation to UNESCO-IHE for providing the best working environment. I remain very grateful and gratified to my family, especially my parents, my wife, my brothers and sister and the two young researchers Adham and Yusuf for their support, patience, understanding and prayers throughout the period of this work.

Curriculum Vitae Moustafa Samir Moussa was born in Cairo, Egypt on August 11th, 1965. He graduated in 1987 as a Civil Engineer from the faculty of Engineering, University of Ain Shams, Cairo. He was awarded his BSc with general grade “very good” and “distinction” for his awarding project. After finishing his military service in 1989 he started his professional career in GOSD, General Organization for Sanitary and Drainage, greater Cairo. Here he was responsible for the sanitary, hydraulic and structural design of Shoubra El Khimma wastewater treatment project. In addition he was appointed as instructor for different training courses in wastewater collection and treatment for the new work orientation training programs. In October 1995, he studied at the International Institute for Infrastructural, Hydraulic and Environmental Engineering (IHE) in Delft (now called UNESCO-IHE Institute for Water Education). In September 1996 he obtained a post-graduate diploma in Sanitary Engineering and was awarded a scholarship from Shell to continue his MSc research. In 1997 he obtained his MSc degree and continued his research through an additional fund from Shell. This research formed the starting point of his PhD research, a joint project of UNESCO-IHE, TU Delft, Shell Global Solutions B.V., Ecco Tannery Holland B.V., Heiploeg B.V. and BTS Senter (an agency of the Dutch Ministry of Economic Affairs). Since Januari 2004 the author works as a researcher/lecturer at UNESCO-IHE.

Nitrification in Saline Industrial Wastewater

Industrial Wastewater Treatment

Nitrification

Saline Solution

Wastewater Bacteria (Wastewater Microbiology)

Saline Solution

Saline Water Conversion II

Model Predictive Control of Wastewater Systems (Advances in Industrial Control)

Retief Saline Solution

Retief Saline Solution

Saline Water Conversion

Wastewater Pathogens

Nitrification in the Activated Sludge Process

Wastewater Microbiology

Industrial Wastewater Management, Treatment, and Disposal, 3e MOP FD-3

Membranes for Industrial Wastewater Recovery and RE-Use

Wastewater Microbiology

Wastewater Management

Wastewater Pathogens

Wastewater Pathogens

Nitrification and Denitrification in the Activated Sludge Process

Wastewater Sludge Processing TQN

Wastewater Sludge Processing

Water and Wastewater Engineering

Water and Wastewater Engineering

Options for Wastewater Management

Natural Wastewater Treatment Systems

Membrane Technology in Water and Wastewater Treatment

Industrial Relations in Schools

Advances in Water and Wastewater Treatment Technology

Nitrification in Saline Industrial Wastewater

Industrial Wastewater Treatment

Nitrification

Saline Solution

Wastewater Bacteria (Wastewater Microbiology)

Saline Solution

Saline Water Conversion II

Model Predictive Control of Wastewater Systems (Advances in Industrial Control)

Retief Saline Solution

Retief Saline Solution

Saline Water Conversion

Wastewater Pathogens

Nitrification in the Activated Sludge Process

Wastewater Microbiology

Industrial Wastewater Management, Treatment, and Disposal, 3e MOP FD-3

Membranes for Industrial Wastewater Recovery and RE-Use

Wastewater Microbiology

Wastewater Management

Wastewater Pathogens

Wastewater Pathogens

Nitrification and Denitrification in the Activated Sludge Process

Wastewater Sludge Processing TQN

Wastewater Sludge Processing

Water and Wastewater Engineering

Water and Wastewater Engineering

Options for Wastewater Management

Natural Wastewater Treatment Systems

Membrane Technology in Water and Wastewater Treatment

Industrial Relations in Schools

Advances in Water and Wastewater Treatment Technology

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