Review

Design and operation of nitrifying trickling filters in

recirculating aquaculture: A review

E.H. Eding a,*, A. Kamstra b, J.A.J. Verreth a, E.A. Huisman a,A. Klapwijk c

a Wageningen University, Department of Animal Sciences, Wageningen Institute of Animal Sciences (WIAS),

Fish Culture and Fisheries, P.O. Box 338, 6700 AH Wageningen, The Netherlandsb Solea BV, Westerduinweg 6, 1976 BV IJmuiden, The Netherlands

c Wageningen University, Department of Agricultural, Environmental and Systems Technology,

Sub-Department Environmental Technology, P.O. Box 8129, 6700 EV Wageningen, The Netherlands

Received 21 September 2005; accepted 21 September 2005

Abstract

This review deals with the main mechanisms and parameters affecting design and performance of trickling filters in aquaculture.

Relationships between nitrification rates and easily accessible process parameters, like bulk phase concentration of TAN, O2,

organic matter (COD), nitrite, temperature, HCO3�, pH and the hydraulic loading of the trickling filter, are discussed in relation to

the design and operation of such filters. Trickling filter design procedures are presented and one of them, a model describing the

nitrification performance of trickling filters by plug-flow characteristics, is discussed in greater detail. Finally, practical aspects in

relation to filter design and operation are presented.

# 2005 Elsevier B.V. All rights reserved.

Keywords: Trickling filter; Recirculation; Nitrification; Biofilm; 1/2-Order kinetics; 0-Order kinetics; Design

www.elsevier.com/locate/aqua-online

Aquacultural Engineering 34 (2006) 234–260

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 235

2. Basic elements of recirculation system design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 236

2.1. Production plan . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 236

2.2. Waste production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 236

2.3. Diurnal variation in waste production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 236

2.4. Water quality limits . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 238

2.5. Suspended solids removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 238

3. Operational aspects of trickling filters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 238

3.1. Aerobic heterotrophic conversion of organic material . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 239

3.2. Nitrification in trickling filters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 239

3.3. Parameters affecting nitrification kinetics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 240

* Corresponding author. Tel.: +31 317 483938; fax: +31 317 483937.

E-mail address: ep.eding@wur.nl (E.H. Eding).

URL: http://www.zod.wau.nl/fcf

0144-8609/$ – see front matter # 2005 Elsevier B.V. All rights reserved.

doi:10.1016/j.aquaeng.2005.09.007

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 235

3.3.1. Effects of TAN and O2 concentrations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 240

3.3.2. Effects of organic matter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 243

3.3.3. Effects of nitrite on biofilm kinetics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 244

3.3.4. Effects of temperature. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 245

3.3.5. Alkalinity and pH limitations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 246

3.3.6. Effects of salinity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 247

3.3.7. Effects of hydraulic loading rate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 248

4. Design concepts for trickling filters in aquaculture. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 248

4.1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 249

4.2. Flow calculations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 250

4.3. Dimensioning/sizing a biofilter. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 251

4.4. Empirical relationships . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 252

4.5. Explanatory relationships. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 253

4.6. A plug-flow model for nitrifying trickling filters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 254

4.7. Some practical aspects of trickling filter design and operation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 256

5. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 257

Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 258

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 258

1. Introduction

Compared with domestic wastewater (Metcalf and

Eddy Inc., 1991; Henze et al., 1997), aquaculture

wastewater has a relatively low concentration of

pollutants (Piedrahita, 2003), and thus, bacterial

biomass yield in treatment systems is also low. To

treat this type of water, bioreactors with a high bacterial

cell residence time are required (Bovendeur, 1989).

Fixed biofilm reactors such as trickling filters show this

typical characteristic.

One of the first studies reporting the use of trickling

filters in aquaculture was presented by Liao and Mayo

(1974). They applied trickling filters for water reuse in

salmonid hatcheries, thereby laying the basis for modern

recirculation technology in aquaculture. Reasons to start

water reuse were due to water shortages, economic gains,

health risks or pollution control (Speece, 1973). When

reusing water from aquaculture operations, oxygen often

becomes the first limiting water quality parameter.

However, oxygen concentrations can easily be restored

by aeration or oxygenation. Moreover, metabolite

concentrations, such as total ammonia nitrogen

(TAN = NH3-N + NH4-N), suspended and dissolved

organic matter and carbon dioxide, also need to be

controlled. As NH3-N is toxic at relatively low levels, the

elimination of TAN is the main goal in designing and

operating a recirculating aquaculture system.

Trickling filters consist of a fixed media bed through

which pre-settled or (micro screen) filtered wastewater

trickles down across the height of the trickling filter.

Because bacterial metabolism requires oxygen, air

needs to be supplied to the biofilm. The aquaculture

wastewater flows downwards over a thin aerobic biofilm

and dissolved substrates diffuse into the biofilm while

others – metabolites – diffuse from the biofilm in the

bulk water. As it trickles down, the water is con-

tinuously oxygenated while the carbon dioxide is

degassed and removed by the ventilated air.

Advantages of trickling filters compared to other

filter types applied in aquaculture are: (1) high process

stability due to constant high oxygen levels; (2) CO2

removal by degassing; (3) water cooling in summer-

time; and (4) simplicity of design, construction,

operation and management. The main disadvantages

of trickling filters are: (1) the relatively low volumetric

removal rates (with consequently large sized biofilters);

(2) biofilm shedding; and (3) risk of clogging when not

properly designed and operated. For certain fish species

additional solids removal is necessary.

In this paper, the basic elements for trickling filter

design will be discussed. Trickling filter design consists

of the following consecutive steps: (1) a production plan

for the determination of the peak feed consumption and

waste production; (2) waste mass balance calculations

linking growth, feed consumption and waste produc-

tion; (3) prediction of the effect of diurnal variation in

waste production; (4) determination of water quality

limits; and (5) selection of a solids removal system.

Operational aspects of trickling filters – which cover the

effects of easy accessible process parameters on biofilm

kinetics – will be subsequently reviewed. Thereafter,

design concepts for trickling filters and some practical

aspects of operative trickling filters will be discussed.

Finally, conclusions and recommendations for future

work are highlighted.

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260236

2. Basic elements of recirculation system design

2.1. Production plan

The design of flow rates and water treatment units

(e.g. a trickling filter), for the control of water quality in

the culture tanks of a recirculation system, is based on

the peak waste production. The peak waste production

coincides with the moment the maximum fish biomass

and feed consumption is realized in a production plan

for a fish species. Based on maximum densities, kg fish/

m2 or kg fish/m3, the minimum culture area and culture

volume are calculated.

2.2. Waste production

Once the maximum feed load is known, the waste

production per kg feed should be calculated. Waste

production data are basic data for flow rate calculations

and for the determination of dimensions of water

treatment units such as trickling filters (Speece, 1973;

Bovendeur et al., 1987; Heinsbroek and Kamstra, 1990;

Losordo and Westers, 1994; Wheaton et al., 1994;

Timmons et al., 2001, 2002; Summerfelt and Vinci,

2004).

A theoretical estimation of the quantities of the

different fish wastes is based on a simplified mass

balance, which links growth, feed intake and waste

production. Heinsbroek (1988) and Heinsbroek and

Kamstra (1990) presented such a mass balance for

European eel (Table 1). Waste production can be

Table 1

Mass balance of dietary feed dry matter (DM), nitrogen (N) and

chemical oxygen demand (COD) in terms of growth and waste

products for European eel (values in g/kg feed)

Feed utilization European eel

Dry

matter

Nitrogen COD

Feed input 900 77 1260

Spilled feed – – –

Fecal loss 315 23 441

Settleable 180 17 252

Non settleable 135 6a 189

Non faecal loss 360 41a 50

Oxygen

consumption

409

(536)b

Growth (gain) 225 13 360

Oxygen consumption of European eel is expressed in g COD/kg feed.

Based on Heinsbroek (1988) and Heinsbroek and Kamstra (1990).a TAN load biofilter is 47 g TAN/kg feed.b CO2 production is approximately 536 g/kg feed.

expressed best per kg feed for defined feeding levels,

growth rates and feed conversion ratios. The quality

and the quantity of the waste are dependent on fish and

feed related aspects (Bovendeur et al., 1987; Heins-

broek, 1988; Einen et al., 1995; Eding and van Weerd,

1999). Since most published data are case specific, each

time a recirculation system is designed commonly

published data on waste production must be validated

for the specific conditions of the designed production

system.

Mass balance calculations for recirculation system

design can also be performed for carbon (C),

phosphate (P) and crude ash. For some of these mass

balances, feed and fish composition data need to be

converted into: (1) nitrogen, g crude protein/

6.25 = g N, generally the factor 6.25 is used but this

value can be lower (Salo-Vaananen and Koivistoinen,

1996); (2) chemical oxygen demand (COD) using the

stoichiometric coefficients for protein, carbohydrates,

fat and ash in COD (1.25, 1.07, 2.9 and 0 g O2

consumption/g protein, carbohydrate, fat and ash,

respectively (Nijhof, 1994a); (3) total oxygen demand

feed (TOD, g O2/kg feed) =P

COD (protein, carbo-

hydrates, fat (g O2/kg feed)) + 4.57 � g Kjeldahl N/kg

feed (g O2/kg feed) (Nijhof, 1994a). The coefficient

4.57 is based upon the stoichiometric coefficient for

oxygen consumption in nitrification, for explanation

see Section 3.2.

2.3. Diurnal variation in waste production

For the design of fish culture systems, knowledge of

the diurnal variation in waste production and O2

consumption is necessary; without correcting for it,

pollutants may temporarily exceed the maximum or

minimum concentration for a certain water quality

parameter (Climit).

To correct for diurnal variation, actual waste

production data (Fig. 1a) are expressed as a ratio of

actual/average waste production (d) (e.g. actual/average

hourly production; Fig. 1b). The values can be

calculated for each water quality parameter separately.

The peak waste production is represented by dmax

(Fig. 1b).

Subsequently, a constant k – as a design value – can

be chosen as a fixed value in the range 1 � k � dmax.

The choice of k determines the fraction of the waste that

temporarily accumulates in the system volume (see

Fig. 1). The larger the chosen value for k, the smaller the

amount of the waste that will accumulate (Fig. 1d).

When k = dmax, no accumulation will take place

(Fig. 1e). The fraction of waste, which temporarily

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 237

Fig. 1. Diurnal variation in TAN production (g TAN/h): (a) variation in relation to feeding; (b) variation expressed as ratio actual/average TAN

production; (c–e) the accumulation of TAN (shaded area) in relation to the removal capacity of the biofilter, (c) filter is designed to remove average

(k = 1) level of TAN production; (d) the filter is designed to remove more than the average but less than the peak production (dmax) of TAN; (e) filter

designed to remove the peak levels of TAN (dmax).

accumulates in the system volume, can be calculated as

described in Eq. (2.1).

PAcc;TAN ¼ PTAN

Z t2

t1

ðd � kÞ dt (2.1)

where PAcc,TAN is part of the TAN production which

temporarily accumulates in the system volume (g);

PTAN the TAN production in g/day;Rðd � kÞ dt is the

fraction of the TAN production which temporarily

accumulates in the system volume over time t.In flow rate calculations, the k value offers the

possibility to correct for diurnal variation in waste

production (Eq. (2.2); Bovendeur et al., 1987;

Heinsbroek and Kamstra, 1990):

Qr ¼���� k � P�

Climit � Cin � PAcc

Vsystem

� ���� (2.2)

where Qr is recirculation flow (m3/day); k the constant

(1 � k � dmax); P the waste (metabolite) production (g/

day); Climit the maximum or minimum for this water

quality parameter (g/m3); Cin the concentration of this

water quality parameter in the influent water (g/m3);

PAcc the part of the production of this water quality

parameter which is temporarily accumulating in the

system (g); Vsystem is the system volume (m3).

From Eq. (2.2), it can be derived that the value of

PAcc/Vsystem (g/m3) becomes larger in intensive systems

with small system volumes and high feed loads per m3

system volume. Consequently, acceptable water quali-

ties can only be maintained by either increasing the flow

rate (Qr) or the system volume (Vsystem). To some extent

Qr and Vsystem are interchangeable (Heinsbroek and

Kamstra, 1990). However, for highly intensive systems

with low system volumes per kg feed/day (1–3 m3), the

possibility to accumulate suspended solids (SS), CO2

and O2 is negligible (large productions/kg feed when

compared with TAN). Therefore, flow calculations for

these wastes are based on k = dmax and PAcc = 0

(Heinsbroek and Kamstra, 1990).

Bovendeur et al. (1987) and Heinsbroek and Kamstra

(1990) predicted for several k-values and a defined

feeding regime the fraction of the TAN production,

which temporarily accumulates in the system volume

(PAcc,TAN). They showed the relationship between the

choice of k, the fraction of TAN temporarily accumu-

lating within the system volume, the flow rate, the

system volume and the bioreactor surface area which

has to be installed in order to control the TAN

concentration in the fish culture units. In their study, the

design value k was used to tune the dynamics in TAN

waste production to the kinetics of TAN removal in the

bioreactor (see Section 4.3).

Wheaton et al. (1994) and Hochheimer and Wheaton

(2000) checked the potential occurrence of lethal TAN

concentrations due to diurnal variation in TAN

production by dividing hourly TAN load by the system

volume and by the hourly filter exchange rates. The

outcome is used to calculate the NH3-N and NH4-N

concentrations for a range of pH values (pH 6–8). The

calculated NH3-N concentration per pH value is

compared with Climit;NH3-N. However, for this method,

it is still necessary to have an estimate for the amount of

TAN temporarily accumulating in the system volume.

Timmons et al. (2001) corrected the TAN production

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260238

(PTAN in kg/day) based on the applied feeding regime.

For a single feeding, the TAN production is assumed to

take place in a 4-h period instead of 1 day. For all other

cases, the time between feedings was used.

Diurnal variation in waste production is strongly

dependent on the applied feeding method (Bovendeur

et al., 1987; Poxton and Loyd, 1987; Heinsbroek and

Kamstra, 1990). Extending the feeding period and

increasing the feeding frequency can significantly

reduce the diurnal variation in waste production (Muir,

1982) and may have important consequences for flow

rate design, water quality fluctuations and treatment

unit dimensions in recirculation systems (Heinsbroek

and Kamstra, 1990).

2.4. Water quality limits

Water quality limits are vital for the design of flow

rates and water treatment units (e.g. trickling filter) as

well as being a controlling factor in determining the

maximum carrying capacity of systems (Colt and

Orwicz, 1991). Unfortunately, there is still a lack of

information that can be used in water treatment unit

design in terms of acceptable lower and upper water

quality limits for fish (Table 2). For example, for

African catfish almost all water quality limits

presented in Table 2 are based on data for other fish

species. The observed water quality values in

recirculation systems for African catfish can signifi-

cantly deviate from the values presented in Table 2

(Bovendeur et al., 1987).

Table 2

Water quality limits, Climit (g/m3)

Water quality

parameter

African

catfisha

European

eelb

Tilapiac Troutc

Temperature (8C) 25–27 23–26 24–30 10–18

O2 (g/m3) 3–8 >6 4–6 6–8

CO2 (g/m3) <25 <25d 40–50 20–30

SS (g/m3) <25 <25 <15e <10e

NH3-N (g/m3) <0.05 <0.05d <0.06 <0.02

TAN (g/m3) <8f <8f <3 <1

NO2-N (g/m3) <0.1 <15g <1 <0.1

NO3-N (g/m3) <100 <100 – –

Chloride (g/m3) – – >200 >200

(-) values unknown.a From Eding and van Weerd (1999).b From Kamstra (1998).c From Timmons et al. (2002).d From Heinsbroek and Kamstra (1990).e From Timmons et al. (2001).f At pH 7.g From Kamstra et al. (1996).

In general, higher tolerance levels for nitrogen

compounds such as TAN and nitrite result in smaller

biofilter dimensions due to higher removal rates at

higher filter influent concentrations (Nijhof and

Klapwijk, 1995). Trout tolerate only low levels of

nitrite; however, this can be controlled when relatively

high TAN removal efficiencies and low TAN influent

concentrations are realized in trickling filters (Nijhof

and Klapwijk, 1995). A high tolerance for nitrate

concentrations results in less make-up water per kg feed

to control the nitrate concentration in the fish culture

units.

The low culture temperature for cold-water species

like trout may affect the size of the trickling filter due to

the lower TAN removal rate per m2 biofilter surface

area. Information on how selecting target values for

water quality in flow calculations and treatment unit

design is presented in Timmons et al. (2002).

2.5. Suspended solids removal

The removal kinetics of substrates in the trickling

filter biofilm is restricted to truly dissolved substances.

However, most aquaculture fish tank effluents contain

high amounts of solids. Suspended solids, when not

sufficiently removed, will clog a trickling filter and

subsequently prevent the complete wetting of the

installed media. To remove as many solids as possible,

drum filters with a mesh size of 30–40 mm are often

installed in eel farms applying trickling filters.

Also submerged filters applied for solids removal

create conditions for high TAN removal rates in

trickling filters, as they also remove part of the

dissolved BOD (Kamstra et al., 1998). Accumulation

of solids on the biofilm may result in an increased

COD (BOD) load and reduced TAN removal rates

(Andersson et al., 1994) while short- and long-term

COD (BOD) loads were also reported to reduce the

TAN removal rate in (trickling filter) biofilms

(Bovendeur, 1989; Zhu and Chen, 2001). It is

therefore of major importance to select a solids

removal unit with a high efficiency in order to

maintain high TAN removal rates per m2 trickling

filter surface area. A comparison of solids removal

units can be found in Chen et al. (1994), Summerfelt

et al. (2001) and Timmons et al. (2002).

3. Operational aspects of trickling filters

This paragraph deals with some selected aspects of

operation and maintenance of trickling filters. The

conversion of organic material, passed through the

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 239

solids removal unit, in the trickling filter is discussed in

general terms, but the nitrification process and the

parameters that affect the process are dealt with in more

detail.

3.1. Aerobic heterotrophic conversion of organicmaterial

Fish tank effluents, having passed a solids removal

unit, still contain fine solid particles and dissolved

organic matter. These substances have a strong negative

effect on the TAN removal rate in trickling filters

(Bovendeur, 1989; Bovendeur et al., 1990) and

submerged biofilters (Zhu and Chen, 2001). Therefore,

the concentrations of these compounds should be

maintained as low as possible. The fraction of this waste

transported towards the trickling filter depends on the

removal efficiency of the installed solids removal unit.

This fraction is either oxidized to CO2 and different

nutrients, assimilated in biomass, passed unchanged

(‘‘inert’’) and/or converted into other organic matter

(Henze, 1997). Bovendeur et al. (1987) observed a

biodegradability of approximately 80% and a COD/dry

matter ratio of 1.4 for organic waste (COD). Using

Eqs. (3.1) and (3.2), a microbiological oxygen demand of

1.42 kg O2 and 1.59 kg O2/kg organic matter removed

(assumed chemical composition C18H19O9N) can be

calculated, respectively (Henze, 1997):

C18H19O9N þ 17:5O2þHþ ! 18CO2

þ 8H2O þ NH4þ ðwithout nitrificationÞ (3.1)

C18H19O9N þ 19:5O2 ! 18CO2þ 9H2O þ Hþ

þNO3� ðwith nitrificationÞ (3.2)

The maximum yield constant for aerobic hetero-

trophic conversion is 0.5–0.7 g COD of bacterial

biomass per g COD of organic matter converted. It is

important to keep the organic waste load for trickling

filters constant and as low as possible because a high

production of heterotrophic bacteria combined with

biofilm detachment (‘‘sloughing’’) may clog a trickling

filter and, unlike submerged filters, backwashing is not

possible. Sloughing can be induced when a biofilm

switches to endogenous respiration due to complete

consumption of all substrates in the outer layer of the

biofilm (Wik, 2003) and is often observed when changes

in waste load occur due to, for example, the harvesting

and grading of fish.

Depending on management and operation of trickling

filters, the average organic weight per m2 trickling filter

media can be as ‘high’ as 88 g/m2 for aquaculture condi-

tions (Shnel et al., 2002). Much lower organic weight per

m2 filter media was reported for a nitrifying trickling filter

in wastewater treatment: 15 and 5 g dry weight for the top

and bottom section, respectively (Person et al., 2002).

3.2. Nitrification in trickling filters

In addition to the heterotrophic conversions in

trickling filters, the nitrification process also needs to be

controlled. Nitrification (Eq. (3.5)) takes place in two

sequential steps: (1) the conversion of ammonium into

nitrite (2) the conversion of nitrite into nitrate (Eqs.

(3.3) and (3.4)):

Based on the stoichiometry of nitrification (Henze,

1997), some basic quantitative relations, which also

serve as operational requirements for trickling filters,

can be calculated:

� th

e consumption of 4.25 g O2/g NH4+-N removed and

4.34 g O2/g NO3-N formed Eq. (3.8) and a consump-

tion of 4.57 g O2/g NO3-N formed (Eq. (3.5)) can be

calculated. The difference, 4.34 g O2/g NO3-N versus

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260240

Fi

w

ox

C

4.57 g O2/g NO3-N, is explained by the fact that

bacteria assimilate inorganic carbon;

� t

he consumption of 1.98 mol HCO3�/mol NH4-N

oxidized which is equal to 1.98 alkalinity equivalents/

mol NH4-N (=8.86 g HCO3�/g NH4-N or 11.74 g so-

dium bicarbonate/g NH4-N oxidized). For a recircu-

lation system, the alkalinity consumption due to

nitrification can be estimated from the nitrogen

fraction of the feed, that is neither retained in the fish

biomass (Table 1) nor discharged by the solids

removal process or the water exchange flow (see

Bisogni and Timmons, 1994). However, in the case of

aquaculture, the consumption of alkalinity due to

nitrification can be lower when, instead of NH4+, NH3

is excreted by fish. In the latter case, about one

alkalinity equivalent per mol NH3 will be consumed;

� t

he production of 0.16 g bacterial biomass

(C5H7NO2) which represents a yield constant of

0.22 g COD/g NH4+-N oxidized;

� t

he production of approximately 1 g NO3-N and

8.33 g H2CO3/g NH4+-N oxidized.

3.3. Parameters affecting nitrification kinetics

In a trickling filter, the rate of removal of substrates

from the recirculating water is determined by their

diffusion rates into the biofilm. Substrates first diffuse

from the bulk liquid into the biofilm through a stagnant

water layer and then into the biofilm. Once in the biofilm,

the substrate is consumed by bacteria in accordance with

the Monod kinetics, although Harremoes (1978) showed

that a 0-order reaction could also be used to describe the

reaction. Removal rates are expressed relative to the filter

surface (g substrate per m2 per day). The nitrification rate

g. 2. TAN removal rate in relation to the concentrations of TAN (CTAN, le

ater temperature is 25 8C and maximum O2 concentration in the bulk wat

ygen and TAN, indicates which substrate limits the TAN removal r�O2=C�TAN < 3:6 TAN is the rate limiting substrate.

in the biofilm depends on many parameters, which can be

divided into both biofilm specific and reactor specific

types (Boller et al., 1994). Following this division, the

subsequent sections will deal with a selected number of

biofilm specific parameters like bulk phase concentra-

tions of TAN, O2, COD, and nitrite, temperature,

alkalinity (HCO3�) and salinity and with hydraulic

surface loading rate (m3 water per m2 cross-sectional

filter area per day) as a reactor specific parameter.

3.3.1. Effects of TAN and O2 concentrationsResearch on fixed biofilm nitrification processes in

trickling filters at aquaculture loading rates (Bovendeur

et al., 1987) showed reaction kinetics corresponding

with the 1/2-order/0-order kinetic model (Fig. 2) for

fixed biofilms as given by Harremoes (1978) and Jansen

and Harremoes (1984). The model is restricted to

removal kinetics for truly dissolved substrates only and

transport of substrates into the biofilm is described by

molecular diffusion. A graphical presentation of this

model for TAN removal rate versus TAN and oxygen

bulk concentration is shown in Fig. 2.

Fig. 2 shows that the transition from 1/2-order

kinetics to 0-order TAN removal kinetics (C�O2=C�TAN),

the related r*, the tolerance of fish for TAN (Climit,TAN)

and the reactor type and mode of operation (CO2) are

important aspects for bioreactor design (Bovendeur and

Klapwijk, 1986; Bovendeur et al., 1987). Based on

possible combinations of TAN and oxygen concentra-

tions, Bovendeur et al. (1987) distinguished four types

of biofilter performances:

1. 0

ft)

er i

ate:

-order removal kinetics in relation to CTAN as a

result of either reaction rate limitation or a con-

and oxygen (CO2, right) for a hypothetical biofilm. Assumed fresh

s 8 g O2/m3 water. The ratio between the two bulk concentrations,

for C�O2=C�TAN < 3:6 O2 is the rate limiting substrate and for

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 241

stant concentration of the rate limiting substrate

(oxygen);

2. 1

/2-order removal kinetics in relation to CTAN (TAN

diffusion limitation), due to a decrease of CTAN

across the filter height as a result of biofilm activity

(CTAN <C�TAN; rTAN ¼ k1

ffiffiffiffiffiffiffiffiffiffiffiCTAN

p);

3. a

lternating 1/2-order and 0-order TAN removal

kinetics in relation to CTAN due to diurnal variation

in TAN concentration;

4. 1

/2-order removal kinetics in relation to CO2(oxygen

diffusion limitation), due to a decrease of CO2across

the filter height as a result of biofilm activity

(CO2<C�O2

; rTAN ¼ k2

ffiffiffiffiffiffiffiffiCO2

p).

Trickling filters designed to be operated at 0-order

TAN removal kinetics and rmax are relatively small but

have no safety factor, so that disturbances in the culture

process directly result in TAN accumulation in the

system volume.

Many commercial trickling filters operate at 1/2-

order TAN removal kinetics only (CTAN <C�TAN) or

alternating 1/2-order and 0-order TAN removal kinetics

(Heinsbroek and Kamstra, 1990; Kamstra et al., 1998).

The safety factor of these filters varies, respectively,

from relatively high to intermediate. For fish species

tolerating only low TAN concentrations, trickling filters

must be designed to operate under 1/2-order TAN

removal kinetics. In well-ventilated trickling filters,

effects of oxygen diffusion limitation on TAN removal

rate are less than 10% (Wik, 1999).

3.3.1.1. Oxygen concentration. Kamstra et al. (1998)

studied 70 full scale trickling filters in eel farms (3

different plastic filter media types) with hydraulic

surface loads varying from 50 to 800 m3/m2/day and

observed oxygen values near saturation level in the

trickling filter effluents. In two pilot scale seawater

trickling filters, Nolting (2000) observed oxygen

saturation levels of 84–91% in the influent and 84–

96% saturation in the effluent. Gujer and Boller (1986)

reported for a tertiary trickling filter (wastewater

treatment) a decrease from 85 to 70% in oxygen

saturation, when the water temperature increased from 5

to 20 8C. Wik (1999) observed oxygen saturation in

bulk water in the middle of their tertiary nitrifying

trickling filter of approximately 80%. Despite these

results, he concluded that the assumption that the bulk

water in trickling filters is saturated with oxygen is

reasonably correct as it results in an error of the

nitrification rate of less than 10%. Near equivalent water

and air temperatures in the trickling filter may result in

stagnant air and poor aeration, thereby reducing

nitrification capacity (Wik, 2003). Trickling filters

can be equipped with ventilation to remedy this effect.

3.3.1.2. TAN concentration. The TAN concentration

at which transition takes place from 1/2-order to 0-order

TAN removal kinetics is an important parameter for

trickling filter design. For a trickling filter biofilm, C�TAN

of 2.2 g TAN/m3 (Table 3d) was observed by Nijhof and

Bovendeur (1990) and 2.4 g TAN/m3 for a submerged

biofilter (Bovendeur and Klapwijk, 1986). Kamstra et al.

(1998) applied a C�TAN ¼ 2 g TAN=m3 for the evaluation

of the performance of trickling filters in commercial eel

farms. Greiner and Timmons (1998) observed no higher

TAN removal rates at trickling filter influent concentra-

tions than above 2.5 g TAN/m3, which indicates that, at

all depths in the filter, TAN concentrations were higher

than C�TAN and – based on their data – a TAN transition

concentration of 2.3 g TAN/m3 could be calculated.

Results reported by van Rijn and Rivera (1990) also

indicate maximum TAN removal rates at an influent

concentration of approximately 2 g TAN/m3, which

results in a lower C�TAN of approximately 1.6 g/m3. A

low C�TAN concentration can be an indication of low

oxygen levels in the biofilter or high COD (BOD) loads

reducing the 0-order TAN removal rate (Bovendeur et al.,

1990). For a seawater (33–34 ppt) trickling filter biofilm,

a C�TAN ¼ 3:0 g TAN=m3 was observed (Table 3d; Nijhof

and Bovendeur, 1990).

3.3.1.3. O2 and TAN as nitrification limiting substratesin trickling filters. The theoretical relationship at

which a change in limiting substrate occurs (Jansen

and Harremoes, 1984; Szwerinski et al., 1986;

Harremoes and Henze, 1997) is given as follows:

C�O2

C�TAN

¼ vO2;TAN:DTAN

DO2

¼ 3:4g O2=m3

g TAN=m3

ðwater temperature 25�CÞ

where:

� C

�O2and C�TAN are bulk water concentrations;

� D

TAN (¼ DNH4) and DO2

are the corresponding

diffusion coefficients (at 25 8C and pure water, see

Gujer and Boller, 1986) and

� v

O2;TAN is the stoichiometric constant for oxygen

consumption relative to ammonium consumption

(4.25 g O2/g TAN; Harremoes and Henze, 1997).

The substrate that penetrates the biofilm the least is the

rate limiting substrate (Szwerinski et al., 1986). In

theory, both substrates in the bulk water are sufficiently

E.H

.E

din

get

al./A

qu

acu

ltura

lE

ng

ineerin

g3

4(2

00

6)

23

4–

26

02

42Table 3

Overview of pilot scale reactors, experimental conditions and kinetic parameters in relation to TAN and NO2-N elimination (PLASTIC supporting medium: Filterpak1-CR50, Mass Transfer Int.,

Heversham, Cumbria, UK, specific surface area 200 m2/m3; void fraction 0.93)

Adaptation conditions

(trickling filter biofilm)

Experimental monitoring

conditions (batch experiments)

rTAN or rNO2-N

(g/m2/day)

0-order or

1/2-order kinetics

Limitating

substrate

(a) Trickling filter HSL = 70–250 m3/m2/day;

temperature:15 8C; Rainbow Trout: 20 kg/m3

Submerged mode, 15 8C, 0–12 g TAN/m3,

pH 7; CO2= 9 g/m3

CTAN = 3 g/m3 rTAN = 0.25 0-order Oxygen

CTAN = 2 g/m3 rTAN = 0.22 1/2-order TAN

CTAN = 1 g/m3 rTAN = 0.14 1/2-order TAN

(a) Trickling filter HSL = 210 m3/m2/day;

temperature 25 8C; pH 7 African catfish:

90–160 kg/m3

Submerged mode; 25 8C, 0–12 g TAN/m,

pH 7; CO2= 6.4 � 0.7 g/m3

CTAN = 9.9 � 6.9 g/m3 rTAN = 0.55 � 0.13 0-order Oxygen

(b) Standard conditions: long-term COD load Submerged mode; full-grown biofilm samples 24 8C,

no COD load; pH 7; CO2= 7 g/m3

5 g COD/m2/day CTAN = 2–10 g/m3 rTAN = 0.6 � 0.1 0-order Oxygen

10 g COD/m2/day CTAN = 2–10 g/m3 rTAN = 0.39 � 0.12 0-order Oxygen

5 g COD/m2/day CTAN = 2–10 g/m3 + CNO2-N = 1–5 g/m3 rTAN = 0.55 � 0.17 0-order Oxygen

10 g COD/m2/day CTAN = 2–10 g/m3 + CNO2-N = 1–5 g/m3 rTAN = 0.33 � 0.09 0-order Oxygen

(b) Standard conditions: long-term COD load Submerged mode; full-grown biofilm samples 24 8C,

no COD load; pH 7 CO2= 7 g/m3 and:

10 g COD/m2/day CTAN = 2–10 g/m3 rNO2-N = 0.37 � 0.13 0-ordera Oxygen

5 g COD/m2/day CNO2-N = 0–8 g/m3 (no TAN supplementation) rNO2-N = 0.69ffiffiffiffiCp� 0:24 1/2-order Nitrite

10 g COD/m2/day CNO2-N = 0–8 g/m3 (no TAN supplementation) rNO2-N = 0.53ffiffiffiffiCp� 0:18 1/2-order Nitrite

5 g COD/m2/day CTAN = 2–10 g/m3 + CNO2-N = 1–5 g/m3 rNO2-N = 0.76 � 0.24 0-ordera Oxygen

10 g COD/m2/day CTAN = 2–10 g/m3 + CNO2-N = 1–5 g/m3 rNO2-N = 0.40 � 0.12 0-ordera Oxygen

(b) Standard conditions Submerged mode; full-grown biofilm samples; 24 8C,

no COD load; CO2= 7 g/m3

pH 8 (CTAN = 2–10 g TAN/m3) rTAN = 0.71 � 0.06 0-order Oxygen

pH 7 (CTAN = 2–10 g TAN/m3) rTAN = 0.56 � 0.09 0-order Alkalinity?

pH 6 (CTAN = 2–10 g TAN/m3) rTAN = 0.20 � 0.10 0-order Alkalinity?

pH 7 (CNO2-N = 0–8 g NO2-N/m3; no TAN) rNO2-N = 0.69ffiffiffiffiCp� 0:24 1/2-order Nitrite

pH 6 (CNO2-N = 3–4 g NO2-N/m3; no TAN) rNO2-N = 0.69ffiffiffiffiCp� 0:24 1/2-order Nitrite

pH 6 (CNO2-N = 3–4 to 8 g NO2-N/m3; no TAN) rNO2-N = circa 1.0 0-orderb (Inhibition2))

(c) Standard conditions:

temperature = 25 8CSubmerged mode; short-term COD load (3–4 h);

temperature = 25 8C; pH 7; CTAN = 2–10 g/m3

CCOD = 0–25 g COD/m2/day ; CO2= 7 g/m3 rTAN = �0.015CCOD + 0.65 0-order Oxygen

CCOD = 1 g COD/m2/day; CO2= 7 g O2/m3 rTAN = 0.63 0-order Oxygen

CCOD = 1 g COD/m2/day; CO2= 3 g O2/m3 rTAN = 0.29 0-order Oxygen

CCOD = 20 g COD/m2/day ; CO2= 7 g O2/m3 rTAN = 0.28 0-order Oxygen

CCOD = 20 g COD/m2/day ; CO2= 3 g O2/m3 rTAN = 0.05 0-order Oxygen

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 243(d

)S

eaw

ater

con

dit

ion

s:H

SL

=1

50

m3/m

2/d

ay,

sali

nit

y=

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–3

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mp

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=2

48C

,

BO

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g/m

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TA

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/m3

Tri

ckli

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ead

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dex

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bio

film

,C

TA

N=

0–

7g

/m3

Sea

wat

er(p

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O2

8.4

g/m

3;

248C

)r T

AN

=0

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11

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TA

N

Sea

wat

er(p

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O2

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AN

=0

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gen

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ater

(pH

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/m3;

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)r T

AN

=0

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rder

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shw

ater

(pH

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/m3;

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)r T

AN

=0

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g=m

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con

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m2;

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00

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per

atu

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load

ing

rate

circ

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m2/d

ay;

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=7

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load

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atch

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(a)

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age

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Nre

moval

val

ues

for

bio

film

sam

ple

sof

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nd

catfi

shpil

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cale

reci

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syst

em(B

oven

deu

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al.,

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oven

deu

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lap

wij

k,1

98

6).

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film

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rate

due

to

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rem

oval

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(Boven

deu

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Sh

ort

-ter

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fect

of

CO

Dlo

adin

gra

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oven

deu

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al.,

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ity

on

TA

Nre

moval

rate

(Nij

ho

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dB

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rese

nce

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po

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itio

nb

yu

nio

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itro

us

acid

.

available at a CO2/CTAN ratio of 3.4 g O2/m3 per g TAN/

m3. Biofilms grown under aquaculture conditions,

showed C�O2=C�TAN ratios close to the predicted

3.4:3.6 (trickling filter, trout, 15 8C) and 3.8 (submerged

filter, African catfish, 25 8C) were reported in Boven-

deur and Klapwijk (1986), 3.6 � 0.8 in Bovendeur et al.

(1987) and 3.8 for a full-grown trickling filter biofilm

(Nijhof and Bovendeur, 1990). Data reported by Nijhof

and Bovendeur (1990) for full-grown seawater biofilm

operated in trickling filter mode indicated a transition

ratio (C�O2=C�TAN) of 2.3.

3.3.2. Effects of organic matterParticulate organic matter may also be problematic

for biofilters in negatively affecting nitrification through

clogging, occupation of the surface area by bacteria

biomass as well as the through the addition of organics

(Wheaton et al., 1994). Particles can easily attach onto

the biofilm surface leading to thicker biofilms; however,

these biofilms do not necessarily result in higher

nitrification rates. Degradation of organic solids in the

biofilm (Henze et al., 1997) may compete with

nitrification thus lowering the overall nitrification

capacity of the trickling filter (Boller et al., 1994;

Andersson et al., 1994).

When high amounts of easily degradable organic

matter are present in a biofilter, the fast growing

heterotrophic bacteria will ‘out-space’ the slow growing

nitrifiers from the aerobic zone in the biofilm as they

compete for oxygen and space (Wik and Breitholz,

1996; Wik, 2003).

Under such circumstances as mentioned, the 1/2-

order/0-order model presented earlier for two substrates

(oxygen and TAN) becomes more complex (Harremoes,

1982; Rauch et al., 1999).

In aquaculture, two types of organic matter loading

rates for trickling filters can be distinguished:

� s

hort-term peak loading rates (3–4 h), often caused by

diurnal variation in waste production due to the

feeding strategies applied;

� s

tructural high organic matter loading rates due to, for

example, low efficiency of the solids removal unit

(Summerfelt et al., 2001), feed utilization differences

among fish species (Heinsbroek, 1988), feed spill

(Nijhof, 1994a) or differences in feed composition

(Cho et al., 1994).

3.3.2.1. Effects of short-term peak organic loadingrates on nitrification and COD removal. Bovendeur

et al. (1990) showed that short-term ‘‘fecal’’ COD

loading (BOD/COD ratio 0.8), under practical aqua-

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260244

culture biofilm loading rates, results in a reduction of

the TAN 0-order removal rate (Table 3c). However,

these effects were usually small because only a tiny

amount of COD is oxidized per g COD removed

(0.065 g O2/m2/day), which has almost no effect on the

oxygen penetration depth in the biofilm. This COD

oxidation rate is almost completely explained by the

reduction in nitrification. The biofilm respiration rate

(nitrification + COD oxidation) was more or less

constant: 2.84 g � 0.22 g/m2/day for TAN 0-order

removal rates in the range of 0.45–0.58 g TAN/m2/

day, COD removal rates of 0–11.4 g/m3 and COD

oxidation rates of 0.268–1.12 g/m2/day thus illustrating

that the removal rate and oxidation rate are not the same.

These observed biofilm respiration rates are relatively

low when compared with the rates of 4–12 g O2/m2/day

presented by Daigger et al. (1994) in three full scale

trickling filters (municipal and industrial wastewater).

3.3.2.2. Effects of structural high organic matterloading rates on nitrification and COD removal. -Long-term COD (BOD) loads on trickling filter biofilms

result in a more pronounced decrease of 0-order TAN

removal rate than short-term COD loads at comparable

heights. This is explained by a higher production of

non-nitrifying material (adsorbed organic matter and

heterotrophic bacteria) resulting in a shorter residence

time of nitrifiers in the aerobic zone (Bovendeur, 1989).

For practical farming conditions, Zhu and Chen

(2001) assume that biofilters will operate at a BOD5/

TAN ratio of approximately 4 (�120 g BOD5 and

30 g TAN production per kg feed). For submerged

filters in series, when loading a biofilm with BOD5/TAN

ratios of 1.76 and 3.52, they observed an average TAN

removal rate of only 0.48 g TAN/m2/day, which is

approximately 30% of the value found under conditions

with only TAN supply. Bovendeur (1989) compared

long-term high COD loaded (BOD/COD = 0.8) trick-

ling filter biofilms (10 g COD/m2/day) with low

(standard) loaded biofilms (5 g COD/m2/day). A pro-

nounced decrease of 30% of 0-order TAN removal rate

was observed (Table 3b) when TAN concentrations of

2–10 g TAN/m3 were applied and of 40% when nitrite-

N and TAN were simultaneously supplied (Table 3b),

thereby showing the difference between removal from

the biofilm and removal from the biofilm and bulk water

simultaneously. Long-term COD loads also resulted in a

significant decrease of the NO2-N removal capacity

(Table 3b).

Using the data of Okey and Albertson (WEF, 2000),

Metcalf and Eddy Inc. (2003) describe the effect of the

wastewater influent BOD/total Kjelhdahl nitrogen ratio

(wastewater treatment) on the nitrification rate

(rN = g N/m2/day):rN = 0.82 (g BOD/g TKN)�0.44.

For trickling filters applied in aquaculture, Heins-

broek and Kamstra (1990) observed a net production of

suspended solids across the filter. However, Parker et al.

(1997) showed that trickling filters can also be operated

as net SS removal units when low BOD and SS influent

loading rates are applied. Differences may be attributed

to the differences in water treatment prior to nitrification

(solids removal and BOD removal) and the hydraulic

loading rate of the trickling filters. The observations of

Parker et al. (1997) give an indication of the operational

conditions of tertiary trickling filters applied in waste-

water treatment and can be compared with data from

trickling filters applied in aquaculture. An average SS

removal rate of 0.2 g SS/m2/day, 0.2 g CBOD5/m2/day

and 1.18 g N/m2/day can be calculated for average

plant influent concentrations of 11.2 g SS/m3, 11.0 g

CBOD5/m3 and 22.8 g NH4-N/m3, respectively. The

observations were made at a temperature between 14

and 22 8C and an average hydraulic surface load of

100 m3/m2 cross-sectional surface area per day (cross

flow media; specific surface area 138 m2/m3/day; filter

height 7.3 m). Maximum TAN removal rates observed

in tertiary trickling filters vary between 1.2 and 2.9 g N/

m2/day (Metcalf and Eddy Inc., 2003) due to high TAN

concentrations and low BOD5/TKN ratios in the filter

influent.

A substantial decrease in 0-order nitrification rate

was observed when high rate nitrifying trickling filters

(wastewater treatment) operating at SS concentrations

<15 mg SS/l (2.6 g N/m2/day) were compared with

operational concentrations >15 mg SS/l (1.8 g N/m2/

day) (Andersson et al., 1994).

In conclusion, high BOD/TAN ratios in trickling

filter effluents and high SS loading rates have a negative

impact on the nitrification rate in trickling filters. BOD

and SS removal to sufficient low concentrations prior to

nitrification in trickling filters may enhance the

nitrification rate in these filters significantly and may

also result in trickling filters operating at a net SS

removal rate.

3.3.3. Effects of nitrite on biofilm kineticsIn literature, relative high concentrations of the

intermediate nitrification product nitrite are often

reported for systems with trickling filters (e.g. Otte

and Rosenthal, 1979; Bovendeur et al., 1987; van Rijn

and Rivera, 1990; Kamstra et al., 1998; Nolting, 2000).

However, Nijhof and Klapwijk (1995) showed that

reported nitrite ‘‘accumulations’’ in recirculation fish

culture systems, in which trickling filters are applied,

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 245

stabilize around an equilibrium concentration at which

all the nitrite produced is oxidized, indicating equal

rates of TAN removal and NO2-N removal.

Nitrite removal differs from TAN removal as the

substrate nitrite is produced within the biofilm layer.

Partial outward diffusion of nitrite from the peripheral

part of the biofilm into the recirculating bulk water will

initially result in elevated nitrite levels in the bulk

water. An equilibrium situation is achieved when the

nitrite concentration in the bulk water is equal to the

nitrite concentration in the peripheral part of the

biofilm. When the TAN concentration increases, a

higher TAN removal rate will result in a higher net

outward diffusion of nitrite into the bulk water. The

equilibrium will be readjusted at a higher nitrite

concentration that prevents the nitrite outward diffu-

sion, eventually leading to equal rates of TAN removal

and NO2-N removal.

Observed nitrite concentrations in trickling filter

effluents in fresh and seawater recirculation systems

could be entirely explained by diffusional transport

mechanisms in combination with biofilm characteristics

(Nijhof and Klapwijk, 1995). By describing TAN and

NO2-N removal rate as a function of the substrate (S)

concentrations, the authors showed that a fixed ratio

NO2-N to TAN concentration can be expected depend-

ing on the ratio of the 1/2-order removal constants:

If rTAN ¼ rNO2-N then affiffiffiffiffiffiffiffiffiffiSTAN

p¼ b

ffiffiffiffiffiffiffiffiffiffiffiffiffiSNO2-N

pand

SNO2=STAN ¼

�a

b

�2

NO2-N concentrations measured in an eel (freshwater,

25 8C) and turbot culture system (salt water, 18 8C)

were proportional to the TAN concentration. The

estimated influent values for a and b were 0.56 and

0.24 in the eel culture system and 0.42 and 0.51 in the

turbot culture system, respectively. For freshwater (eel

culture), a much higher substrate ratio (SNO2/

STAN = 4.0) was observed when compared to that

observed for seawater (SNO2/STAN = 0.4). Nolting

(2000) found values of the same order of magnitude

at 24–26 8C: NO2-N/NH4-N ratios of 1.3–3.1 at low

salinity (16 ppt) and ratios of 0.6–0.8 at high salinity

(30 ppt).

The clear proportionality between NO2-N and TAN

concentrations, as observed by Nijhof and Klapwijk

(1995), could not be confirmed for commercial farms

(Kamstra et al., 1998). An average NO2-N/TAN ratio of

1.3 � 1.6 was found, the ratios ranging from 0.2 to 7.7,

showing large variations between recirculation systems.

Kamstra et al. (1998) conclude that nitrite oxidation

capacity in biofilms seems to be variable and sensitive to

environmental disturbances.

During their experiments, Nijhof and Klapwijk

(1995) never observed a net nitrite production across

the trickling filter. Although van Rijn and Rivera (1990)

observed net nitrite production, their observations are

probably based on measurements during peak ammonia

production, when no steady state of ammonia and nitrite

concentrations was reached. Otte and Rosenthal (1979)

also presented a relatively constant NO2-N/TAN ratio in

a closed brackish water system with a long-term

operating trickling film.

Developing trickling filter biofilms show a delay in

NO2-N removal capacity when compared with the final

TAN removal capacity. However, NO2-N removal

capacity exceeds TAN removal capacity when the

biofilm develops into a full-grown biofilm (Bovendeur,

1989). Full-grown biofilms may accommodate consider-

able nitrite removal capacity (maximum � 1.7 g NO2-N/

m2/day; Table 3d) when compared to the 0-order TAN

removal capacity (0.6 g TAN/m2/day; Table 3d). These

results were obtained when biofilms were supplied with

one substrate only (nitrite or TAN, respectively;

Table 3d). In cases of simultaneous TAN and nitrite

load, nitrite removal rate shifted from 1/2-order removal

rate to 0-order removal rate at lower nitrite removal rates

(Table 3d).

Nijhof and Klapwijk (1995) observed that the ratios

of TAN removal to NO2-N removal vary between

biofilms but the observed relatively high nitrite removal

rates were neither a result of a long-term operating

biofilm nor of salinity. High levels of nitrite were not

caused by an inhibition of the nitrification process.

In contrast to the TAN oxidation capacity, the nitrite

oxidation capacity proved homogeneously distributed

across the height of the trickling filter in the eel culture

system (Nijhof and Klapwijk, 1995).

The results of Nijhof and Klapwijk (1995) indicate

that low nitrite filter effluent concentrations depend on:

(1) low TAN concentrations in the filter influent; and (2)

the ratios TAN removal rate to NO2-N removal rate in

the biofilter.

3.3.4. Effects of temperatureIn a study, Boller and Gudjer (1986) corrected

nitrification rates for temperature in nitrifying tertiary

trickling filters using the following equation:

rTAN,10 8C = rTAN,T exp[kT(10 � T)], where kT = 0.044

8C�1, T, ambient water temperature (8C).

Bovendeur et al. (1987), using from van’t Hoff-

Arrhenius developed equation rT = r20u(T � 20), com-

pared the 0-order TAN removal rate of two indepen-

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260246

dently operated trickling filters at 15 8C (rTAN = 0.25 g -

TAN/m2/day) and 25 8C (rTAN = 0.55 g TAN/m2/day)

(Table 3a). They concluded that the difference in rTAN

was completely attributed to the observed temperature

activity coefficient u (observed u = 1.08). However, the

temperature activity coefficient for trickling filters

shows a range from 1.02 to 1.08 (Metcalf and Eddy Inc.,

1991), leading to a wide range in TAN removal rates.

Okey and Albertson (1989a,b) re-interpreted studies

on tertiary trickling filters for temperature effects and

oxygen limitation. They concluded that the concentra-

tion and diffusivity of oxygen control the nitrification

rate in cases where the TAN concentration itself is not a

limiting factor. Zhu and Chen (2002) report similar

findings where the impact of temperature on nitrifica-

tion rate in a fixed submerged biofilm is less significant

than predicted by the van’t Hoff-Arrhenius equation. At

a temperature of 20 8C, they report an increase in the

nitrification rate of 1.108% for 1 8C increment under

oxygen limitation conditions and 4.275% under TAN

limitation conditions.

For rotating biological contactors, Wortman and

Wheaton (1991) reported a linear relationship between

TAN removal rate and temperature (for the temperature

range 7–35 8C): v(g TAN/m3 filter/day) = 140 +

8.5 T(8C). For complete nitrification, a linear relation-

ship of v(gNO3-N/m3 filter/day) = 63 + 9.9 T(8C) was

reported. They concluded that ammonia oxidizing

bacteria and nitrite oxidizing bacteria showed a similar

sensitivity to temperature since the slope of both

equations was not significantly different. The impact of

temperature on 0-order TAN removal rate studied by

Zhu and Chen (2002) and Wortman and Wheaton

(1991) is based on a two-component substrate model

(TAN and O2) and simultaneous effects of BOD and

suspended solid loads were not incorporated in their

studies. Moreover, tertiary trickling filters are operated

at low BOD loading rates and temperature effect on

nitrification rate is mainly masked in the concentration

range for which TAN is not rate limiting. Many trickling

filters in aquaculture are however operated at 0-order

and 1/2-order TAN removal rates or at 1/2-order rate

only which implicates that temperature effects should

not be neglected. Although recently, more research is

being devoted to the effects of temperature on

nitrification; these results seem to be contradictory.

3.3.5. Alkalinity and pH limitationsAccording to Eq. (3.8), nitrification requires 2 mol

HCO3� per mol NH4

+ oxidized and and 1 mol HCO3�

in case of oxidizing 1 mol NH3, which equals two and

one alkalinity equivalent per mol substrate oxidized,

respectively. This will lower the pH in the bulk water as

well as in the trickling filter biofilm, and will result in

inhibition of the nitrification process.

3.3.5.1. Effects of alkalinity on TAN removal rate. -For tertiary trickling filters (wastewater treatment),

Gujer and Boller (1986) observed a 100% reduction in

the nitrification rate when the bulk water alkalinity

dropped from 2 mequiv/l (pH 7.7) to 0.2 mequiv/l (pH

6.2) and showed that, for complete nitrification,

alkalinity should not drop below 1.5–2.0 mequiv/l.

Szwerinski et al. (1986) verified the theoretical

predicted pH effect on a 400 mm thick reaction rate

limited nitrifying biofilm using the theory of outward

diffusion. At 2–2.5 mequiv/l alkalinity, they observed a

pH difference between the bulk phase and the biofilm

but no effect on 0-order TAN removal rate. Decreasing

the bulk water alkalinity to 0.7 mequiv/l (�pH 6.7)

resulted in a decrease of the TAN removal rate. A strong

drop in the biofilm pH of �6.5 to �5.8 was observed at

a bulk water alkalinity below 0.7 mequiv/l. At biofilm

pH 5.7, the pH of the bulk water remained more or less

constant due to hydrogen ion toxicity. Biesterveld et al.

(2003) suggested that, in addition to the alkalinity

destruction by the nitrification process, a minimum

level of carbonate alkalinity (0.9 mequiv/l) must be

present to cover inorganic carbon requirements of the

ammonia oxidizers. The effect seemed to be indepen-

dent of the pH in the range of 7–8.

Alkalinity rate limitation and pH inhibition in the

biofilm can also be predicted on the basis of the ratio at

which one of the substrates (HCO3/NH3-N and HCO3/

O2) becomes diffusion limiting (see Section 3.3.1 and

Szwerinski et al., 1986).

3.3.5.2. Effects of pH on TAN removal rate. In a

literature overview, Wheaton et al. (1994) presented the

pH optima for nitrosomonas (pH 6–9) and nitrobacter

(pH 6.3–9.4) and mentioned that the operational range

is probably from pH 5 to 10 provided that the biofilm

can adapt slowly. However, complete cessation of

nitrification at a pH of 5.5 was also reported.

Nitrifying bacteria in biofilms, however, ‘‘experi-

ence’’ a pH which is lower than the bulk water due to

mass transfer resistance (Siegrist and Gujer, 1987;

Szwerinski et al., 1986).

Bovendeur (1989) observed a pH induced inhibition

(short-term measurement) of 0-order TAN removal rate

for fixed film biofilms with an average TAN removal

rate of 0.71 g/m2/day at pH 8) grown in trickling filter

mode at aquaculture waste loading rates (Table 3b). The

data can be used to predict the reduction in nitrification

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 247

for the pH range 6–8 using the polynomial y =

0.148x2 � 2.43x + 9.972 (y, fractional reduction in

nitrification; x, pH (range 6–8). A reduction in 0-order

TAN removal rate of approximately 70% can be

calculated for pH 6.

Kruner and Rosenthal (1983) showed that nitrifica-

tion in trickling filters, under the salinity of 14–16 ppt,

was reduced to almost zero at pH 5.6. Comparable

reductions due to pH were reported in other studies

using rotating contact filters and other fixed film

biofilters (Boller et al., 1994). In these filters,

characterized by high TAN removal rates, nitrification

of TAN stops completely at a bulk fluid pH of 6.5–6.7.

This fits with the observations of Szwerinski et al.

(1986) where pH levels below 6.7 in the bulk fluid

resulted in a drop of the biofilm pH close to the support

media of 5.7.

Based on nitrification results from a fluidized bed

biofilm reactor operated at pH 4.5, pure oxygen supply

and temperature 25 8C, Tarre et al. (2004) conclude that

the inhibition of nitrification at low pH as reported in

literature is probably highly overestimated. Instead of

pH, CO2-limitation due to excessive CO2-degassing is

probably the reason why high nitrification rates at low

pH have not been observed earlier (Tarre and Green,

2004; Green et al., 2002).

In the Netherlands, many farms operate trickling

filters with a pH in the bulk fluid below pH 7. Despite

the long-term adaptation period of nitrifiers to low pH

values, high TAN and nitrite concentrations are still

observed indicating that adaptation will not result in the

re-establishment of removal rates in the optimal pH

range. As trickling filters are excellent CO2-degassers,

CO2-limitation might be a reason for low TAN removal

rates at low pH (Green et al., 2002).

Therefore, and in accordance with the recommenda-

tions of the U.S. EPA (2000), the effect of pH below the

neutral range, if anticipated, should not be ignored when

dimensioning trickling filters.

3.3.5.3. Effects of pH on NO2-N removal rate. Full-

grown biofilms maintain a considerable nitrite removal

capacity at low pH (Bovendeur, 1989). When only

nitrite was supplied as substrate, 1/2-order nitrite

removal rates up to �1.7 g NO2-N/m2/day were

observed for nitrite concentrations of 0–8 g NO2-N

(g/m3) at pH 7 (Table 3b). At pH 6, a transition of 1/2-

order nitrite removal kinetics to apparent 0-order

nitrite removal kinetics was observed for nitrite

concentrations exceeding 3–4 g NO2-N/m3, resulting

in a NO2-N removal rate of approximately 1.0 g/m2/day

(Table 3b). The reduction in nitrite removal capacity at a

lower pH was thought to be caused by inhibition of

unionized nitrous acid.

3.3.6. Effects of salinityFor seawater trickling filter biofilms, a slower start up

to ‘full-grown biofilm stage’ and a lower TAN removal

rate was observed when compared with freshwater

biofilms (Nijhof and Bovendeur, 1990). A considerably

longer start-up period (�170 days) was needed for

seawater trickling filter biofilm to reach the ‘full-grown

stage’ when compared to a freshwater biofilm (�130

days). ‘Full-grown’ seawater trickling filter biofilm

samples from a commercial seawater eel farm also

showed considerably lower 0-order TAN removal rates

(0.28 g/m2/day) when compared to freshwater biofilm

samples (0.69 g/m2/day) (Table 3d). Similar to fresh-

water biofilms, the TAN removal rate for seawater

biofilms could also be described using the 1/2-order/0-

order model (Table 3d). However, it is not clear whether

these lower observed TAN removal rates for seawater

when compared with freshwater are: (1) a specific result

of this research or (2) whether they should be seen as a

general effect of seawater on the nitrification rate.

For instance, Nijhof (1994b) reported a maximum

TAN removal rate of �0.9 g TAN/m2/day for experi-

ments with batch wise examination of a 60 l seawater

(about 34 ppt) trickling filter. These removal rates are

similar to TAN removal rates observed for biofilms

from the top of a freshwater trickling filter (Nijhof,

1995). Nijhof and Klapwijk (1995) also reported a high

0-order TAN removal rate of approximately 0.7 g TAN/

m2/day for a ‘full-grown’ biofilm sample in the upper

part of the trickling filter described in Nijhof (1994b).

Nolting (2000) observed NH4-N removal rates of 0.06–

0.24 g NH4-N/m2/day for pilot scale trickling filters

operating at 16 ppt seawater and loading rates of 0.1–

0.4 g NH4-N/m2/day. After adapting the trickling filters

from 16 to 30 ppt seawater, the highest removal rate

observed was 0.66 g NH4-N/m2/day at a loading rate of

approximately 2.7 g NH4-N/m2/day. However, it is

unclear whether the biofilm in this study had reached

the ‘full-grown stage’.

When comparing the observed seawater TAN

removal rates with freshwater TAN removal rates

(Nijhof, 1994b; Kamstra et al., 1998) as function of

TAN loading rates, seawater TAN removal rates in the

TAN loading range of �0.3 and �0.45 g TAN/m2/day

were not lower than those observed for freshwater.

However, the lower oxygen concentration in saturated

seawater when compared to freshwater may result in

lower maximum TAN removal rates for the concentra-

tion range where TAN is not limiting.

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260248

Further research under standardized conditions should

be performed in order to be able to draw conclusions with

respect to possible differences (if any) in TAN removal

capacity between fresh- and seawater biofilms.

Nitrite removal rate in seawater is significantly

slower to develop than in freshwater. A considerably

larger nitrite accumulation was observed when com-

pared to freshwater during the first months of the start-

up phase due to the slow development of the nitrite

oxidation capacity (Nijhof and Bovendeur, 1990).

Batch experiments with a several year old biofilm

sample (from the upper part of a seawater (turbot culture)

or freshwater (eel culture) pilot scale trickling filter),

showed rNO2-N > rTAN for seawater and rNO2-N < rTAN

for freshwater (Nijhof and Klapwijk, 1995). It is not clear

if this is a structural difference or caused by differences in

culture conditions between eel and turbot. The reported

homogenous distribution of rNO2-N across the height of a

freshwater trickling filter (Nijhof and Klapwijk, 1995)

was not determined for seawater trickling filters.

To overcome long start-up periods in seawater

recirculation systems, Nijhof and Bovendeur (1990)

showed that two freshwater trickling filter biofilms with

a 0-order TAN removal rate of approximately 0.3 g/m2/

day could be adapted after a prompt switch from fresh to

seawater (17 or 34 ppt) at day 1. After an adaptive

period of approximately 10 days, 0-order TAN removal

rates in both seawater biofilms were comparable to the

initial freshwater TAN removal rate. However, nitrite

oxidation capacity of a freshwater biofilm sample was

much more vulnerable to elevated salinities than the

corresponding ammonia oxidation capacity and a

period of adaptation to intermediate salinity was

strongly recommended (Nijhof and Bovendeur, 1990).

3.3.7. Effects of hydraulic loading rateIn literature, minimum and maximum hydraulic

surface loading rates (HSL; m3/m2 filter cross-section/

day) are reported for trickling filters (Wheaton et al.,

1994). The upper and lower limits for HSL vary with

specific surface area and media type. Head loss or

removal of bacteria from the plastic media limits the

increase of hydraulic surface loading (Wheaton et al.,

1994). A minimum HSL is necessary to keep the

complete filter surface area wet (Boller and Gudjer, 1986;

Wheaton et al., 1994) and may be needed to control the

concentration of grazing organisms in a trickling filter

(Boller and Gudjer, 1986). Minimum hydraulic loading

rates reported for trickling filters are 32–55 m3/m2/day

for random packed plastic pall rings (Roberts, 1985), and

29 m3/m2/day for randomly packed Norton Actifil1

media (Grady and Lim, 1980) (both quoted in Wheaton

et al., 1994). Boller and Gudjer (1986) reported a positive

effect of hydraulic surface loads 72 m3/m2/day on the

control of biofilm grazers in trickling filters (specific

surface area 230 m2/m3). Bovendeur et al. (1987) found

100–200 m3/m2/day suitable as HSL for random filter

media (Filterpac CR50, specific surface area 200 m2/m3,

void fraction 0.93). Maximum hydraulic loading rates

reported are 72–188 m3/m2/day for plastic pall rings

(Roberts, 1985), 234–350 m3/m2/day for Dow surfpac1

(Grady and Lim, 1980) (both quoted in Wheaton et al.,

1994). Nijhof (1995) tested hydraulic surface loads of

75–300 m3/m2/day for Filterpac CR50. Kamstra et al.

(1998) evaluated the performance of commercial

trickling filters in eel farms and reported a minimum

observed HSL of 50 m3/m2/day (Filterpac CR50, random

media, 200 m2/m3) and a maximum observed HSL of

800 m3/m2/day (Munters C10.12, 234 m2/m3, cross flow

media).

Nijhof (1995) studied the effect of three hydraulic

surface loading rates (75, 150 and 300 m2/m2/day) on

the nitrification rate in trickling filters (Filterpac CR50,

random media; freshwater, eel culture; 25 8C, filter bed

height 1.5 m). He observed a clear effect of HSL on the

half order rate constant. The positive effect of an

increased HSL on nitrification rate can be explained by

improved wetting, increased TAN loading rates

(Kamstra et al., 1998), prevention of a non-continuous

biofilm development in the lower part of the trickling

filter (Boller and Gudjer, 1986), or by increased

availability of bicarbonate in the lower part of the

trickling filter (Siegrist and Gujer, 1987). Greiner and

Timmons (1998) tested HSL’s of 469–1231 m3/m2/day

at a temperature of 26.4 8C using 5.1 cm Norpak media

(NSW Corporation, Roanoke, VA) in pilot sized

biofilters and did not observe any effect of hydraulic

loading rates on the nitrification rate. The nitrification

rates observed by Greiner and Timmons (1998) are even

higher than observed in other studies in aquaculture,

even when TAN was the only substrate in combination

with oxygen (Zhu and Chen, 2002). Greiner and

Timmons (1998) concluded that these findings, in

combination with the applied higher HSL range and

concomitantly higher nitrification rates (0.92–3.92 g/

m2/day), suggest that there is a limit to the effect of HSL

on nitrification rate.

4. Design concepts for trickling filters in

aquaculture

This section presents first some commercial applica-

tions of trickling filters in recirculation systems.

Subsequently, design concepts and design methods are

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 249

Table 4

Combination of solids removal, trickling filters and types of fish tanks installed in commercial recirculation systems

Fish species Fish tanks Solids removal unit Reference

African catfish Rectangular tanks Lamella sedimentation or triangle filter Verreth and Eding (1993)

European eel Circular tanks Drum filter or triangle filter Kamstra et al. (1998)

European eel Rectangular tanks Submerged up flow filter

Tilapia Dual drain tank technology Drum filter Twarowska et al. (1997)

Losordo et al. (2000)

Tilapia Circular tanks,

twice-daily drainage

of solids in the tank

Drum filter Shnel et al. (2002)

Salmon smolts Circular tanks Drum filter + up flow filter De Bondt (personal communication)

Turbot Octagonal tanks Drum filter Eding and Kamstra (2002)

reviewed and flow calculations and biofilter dimension-

ing for TAN-control are discussed. Empirical and

explanatory relationships for the determination of

TAN-removal are presented, with special attention to a

plug-flow model, which was validated for a large number

of commercial trickling filters. Finally, practical aspects

in relation to filter design and operation are discussed.

Fig. 3. Left: a typical recirculation system in Dutch African catfish f

4.1. Introduction

Trickling filters are widely applied in recirculation

systems (Table 4; Fig. 3). Depending on the sensitivity

of farmed species for particulate matter, trickling filters

are used in combination with one of the following solids

removal units: sedimentation (Fig. 3); a drum filter or

arms. Right: a typical recirculation system in Dutch eel farms.

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260250

disc filter; a submerged up flow filter; or with a drum

filter and a submerged up flow filter (Fig. 3).

Several papers have described the design of trickling

filters for aquaculture use. Liao and Mayo (1974)

described the relationship between TAN removal rate,

filter media retention time, TAN loading rate and water

temperature. A detailed design procedure incorporating

the findings of Liao and Mayo (1972, 1974), was

presented by Wheaton (1977).

Bovendeur et al. (1987) incorporated 1/2-order and

0-order TAN removal kinetics in a design concept for

recirculation systems. This concept was used by

Heinsbroek and Kamstra (1990) for the evaluation of

four commercial recirculation systems and one pilot

installation for the culture of European eel.

Nijhof (1995) used a plug-flow model to predict the

TAN removal for a pilot scale trickling filter emphasiz-

ing plug-flow characteristics, 1/2-order/0-order TAN

removal kinetics, the TAN influent concentration,

hydraulic surface load and the observed stratification

in TAN oxidation capacity. Kamstra et al. (1998)

validated the plug-flow model of Nijhof (1995) for a

range of full-scale trickling filters at 14 commercial

eel farms (70 commercial trickling filters) in the

Netherlands.

Design methods for trickling filters are presented by

Wheaton et al. (1994), Hochheimer and Wheaton

(2000) and Timmons et al. (2001, 2002). A spreadsheet

procedure for flow and biofilter (including trickling

filters) sizing was developed by Losordo et al. (2000).

Hochheimer (1990) developed a mathematical model of

an aquacultural trickling filter. The computer model was

validated with data from six laboratory scale filters.

4.2. Flow calculations

The procedure for flow calculations should initially

focus on the maximum feeding rate (kg feed/day),

maximum biomass and culture volume (see Section 2.1)

and the waste production per kg feed (see Section 2.2)

(Wheaton et al., 1994; Losordo et al., 2000; Timmons

et al., 2002).

For flow calculations, a mass balance analysis is

needed for the different nutrients relevant to the target

water quality parameter, thereby assuming a steady state

of effluent concentration per water quality parameter for

the control volume. Steady state conditions imply that

there is no accumulation (dC/dt = 0). In these calcula-

tions, mixed tank conditions are assumed (Metcalf and

Eddy Inc., 2003; Losordo and Westers, 1994; Timmons

et al., 2002; Vinci et al., 2004; Summerfelt and Vinci,

2004a; Summerfelt and Vinci, 2004b).

For flow rate calculations and biofilter design, the

concept presented by Liao and Mayo (1972, 1974) is often

cited. They described the concentration of a metabolite at

the outlet of a fish culture tank in a recirculation system as

a proportion to the concentration of the same metabolite in

a system without recirculation. That proportion coeffi-

cient C is therefore a measure for the accumulation of the

metabolite due to recycling (see Eq. (4.1)). Other authors

used this metabolite accumulation factor to estimate the

concentration of different metabolites at the outlet of a

culture tank (Timmons et al., 2001; Summerfelt et al.,

2001) (Eq. (4.2)). Using Eq. (4.2), it can be deduced that

the desired TAN concentration in the fish tank effluent

(Wasteout) is determined by:

� t

he accumulation factor which is based on the fraction

of the water flow that is reused (R) and the TAN

removal efficiency (treatment efficiency (TE) is the

decimal fraction of a metabolite removed by a single

pass though the treatment unit);

� t

he TAN production rate (PTAN);

� t

he TAN concentration in the make-up water

(Wastenew);

� a

nd the flow rate (Q) controlling the TAN accumula-

tion in the fish tank effluent.

Many recirculation systems are operated at a water

recycling percentage of 96% or more (R 0.96). In

these systems, assuming the make-up water contains

negligible amounts of TAN, the TAN accumulation

depends mainly upon the treatment efficiency across the

biofilter (Eq. (4.3); Timmons et al., 2002). The TAN

effluent concentration of a trickling filter (Ctreatment,out)

in such a recirculation system is based on the desired

fish tank effluent concentration (Ctreatment,in) and the

treatment efficiency (Eq. (4.4); Timmons et al., 2002).

Subsequently, the flow rate for TAN control across the

fish culture units can be calculated with Eq. (4.5):

C ¼ 1

1� Rþ R� TE(4.1)

Wasteout ¼�

1

1� Rþ ðR� TEÞ

����

Pwaste

Qr

�

þ ð1� RÞ � ðWastenewÞ�

(4.2)

CTAN;out ¼�

1

TE

���

PTAN

Qr

�(4.3)

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 251

Ctreatment;out

¼ Ctreatment;in þ TEðCtreatment;best � Ctreatment;inÞ(4.4)

Qr ¼PTAN

TE� CTAN;out

¼ PTAN

CTAN;out � CTAN;in(4.5)

where C is allowable waste concentration in the fish

tank effluent(g/m3)/single pass waste concentration (g/

m3); R the fraction of the water flow reused (fraction);

TE the treatment efficiency (decimal fraction); Wasteout

the waste (metabolite) concentration in the fish tank

effluent (g/m3); Pwaste the waste production of a meta-

bolite (g/day); Qr the water flow, for TAN the water flow

recirculated across the biofilter (m3/day); Wastenew the

concentration of a metabolite in the make-up water (g/

m3); CTAN,out the TAN concentration in the fish tank

effluent (g/m3); CTAN,in the filter effluent concentration

and fish tank influent concentration; Ctreatment,best,TAN is

0 (Timmons et al., 2002); PTAN the production of TAN

(g/day); CTAN,in is the TAN concentration of the fish

tank influent (g/m3).

Some remarks have to be made in relation to the flow

rate calculation for TAN control:

� H

igher flow rates may be needed in order to meet the

hydraulic requirements for trickling filter operation.

� T

AN concentrations in the bulk water higher than

C�TAN (concentration at which transition takes place

form 1/2-order TAN removal kinetics to 0-order

kinetics) will not result in higher TAN removal rates.

� T

he design procedure followed by Losordo et al.

(2000) offers the possibility to reduce the TAN load

(PTAN) for flow calculation and biofilter dimensioning

due to passive nitrification; that is, nitrification

outside the biofilter. However, the surface area of

the fish tanks and piping is relatively small in relation

to the surface area installed in the biofilter and is

loaded with easily degradable organic matter which

would result in relatively low TAN removal rates per

m2 in intensive production systems. High TAN

removal rates are only expected in the piping towards

the fish tanks (high TAN load due to high hydraulic

load and low BOD loading rate).

� T

he flow requirements for the control of each water

quality parameter are calculated in order to dete-

rmine which flow will become the controlling flow

(Heinsbroek and Kamstra, 1990; Losordo and

Westers, 1994; Eding and van Weerd, 1999; Timmons

et al., 2002).

4.3. Dimensioning/sizing a biofilter

For dimensioning or sizing a trickling filter, only

limited information is available. In practice, TAN

removal efficiency is often empirically determined for a

fixed set of successful conditions such as fish species,

feed load, filter height, filter media type, hydraulic

surface load, suspended solids unit and TAN influent

concentration. The set of conditions is transferred to a

‘new’ design and often functions without any problem.

However, when water quality control problems occur,

they are then related to the introduction of either new

filter media, a change in feed composition, water

treatment units being replaced in favour of ‘better and

cheaper’ models or new tank designs. The effect of

some of these changes could have been predicted.

When the TAN removal efficiency for a certain

trickling filter influent concentration is known, it is

based on data for a fixed filter height, media type,

hydraulic surface load, TAN removal rate and

temperature. The required total nitrification surface

area (A, m2; Eq. (4.6)) is calculated from the trickling

filter TAN load (PTAN load,trickling filter, g/day) and the

estimated nitrification rate (rTAN, g TAN/m2/day). The

bioreactor volume (Vtrickling filter, m3; Eq. (4.7)) is a

function of the total filter surface area (A, m2) and the

specific surface area (a in m2/m3 biofilter media) of the

filter media. The shape of the reactor (Eq. (4.8)–(4.10)

depends on the hydraulic surface load (HSL, m3/m2/

day) (Losordo et al., 2000; Wheaton et al., 1994).

Atrickling filter ðm2Þ ¼ PTAN load;trickling filter ðg=dayÞrTAN ðg=m2=dayÞ (4.6)

Vtrickling filter ðm3Þ ¼ Atrickling filter ðm2Þa ðm2=m3 biofilter mediaÞ (4.7)

Scross-sectional area ðm2Þ ¼ Qtrickling filter ðm3=dayÞHSL ðm3=m2=dayÞ (4.8)

Ddiameter filter ðmÞ ¼ 2�

ffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi�Scross-sectional area ðm2Þ

3:1416

�s

(4.9)

Hheight ðmÞ ¼Vtrickling filter ðm3Þ

Scross-sectional area ðm2Þ (4.10)

However, for optimization of trickling filter design, it

is necessary to establish relationships between opera-

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260252

tional parameters and TAN removal rate (Gujer and

Boller, 1986) in order to predict the TAN removal

efficiency of biofilters and find the best set of parameter

values. For the determination of the TAN removal rate

in trickling filters, both empirical relationships (Liao

and Mayo, 1974) and explanatory relationships based

on nitrification kinetics can be used (Bovendeur et al.,

1987; Heinsbroek and Kamstra, 1990; Nijhof, 1995;

Kamstra et al., 1998).

4.4. Empirical relationships

Liao and Mayo (1974) observed that TAN removal

rate (NAR, g TAN/m2/day) is a function of the TAN

loading rate (AL, g TAN/m2/day) and media retention

time (tm = Vmedia (m3)/void fraction/flow rate (m3/h):

NAR = 0.96ALtm). This equation was rearranged in: NAR/

AL = EA (filter efficiency) = 0.96 tm. This approach

Box 1. Trickling filter design procedure based on Liao

At the start of the design procedure, the fraction (R) of

known.

Step 1. Determination of water flow (m3/day) needed fo

Determination of allowable TAN concentration in the fis

filter design, the single pass concentration of TAN has

Step 2. Determine the ammonia accumulation factor (C

C ¼ Allowable ammonia concentration ðClimit;TAN ðg=m3ÞSingle pass ammonia concentration ðCTAN ðg=m3ÞÞ

The single pass ammonia concentration is the ammonia

a single pass system (=TAN production (g/day)/water fl

Step 3. Determine the filter efficiency (E):

E ¼ 1þ CR � C

CR

E: filter efficiency (decimal fraction); C: ammonia accum

Step 4. Calculate the total ammonia load filter (g TAN/

Total ammonia load ¼ ðTAN productionÞ ðCÞStep 5. Calculate filter retention needed to achieve am

tm ¼E

9:8ðT Þ � 21:7

E: filter efficiency (%); tm: media retention time (h); T:

Step 6. Calculate filter volume:

Filter volume ðm3Þ ¼ ðflow rate ðm3=dayÞ � ðretention tim

Step 7. Filter surface area (A, m2):

A ðm2Þ ¼ ðfilter volume ðm3ÞÞðspecific surface area filter m

Step 8. Check if the TAN load is less then 0.977 g/m2/d

Step 9. Determine the filter dimensions.

works only within a given range of conditions, e.g.

temperature (10–15 8C); hydraulic surface loading

(86.4–147 m3/m2/day per cross-sectional surface area);

pH (7.5–8); filter media: 3.5 in. Koch rings; media

retention time 0.206–0.46 h; TAN concentration � 1 g/

m3; and a maximum TAN loading rate of 0.977 g/m2/day

due to organic loading (Wheaton, 1977). The maximum

TAN loading rate is likely to be due to the operational

conditions of their experiments, in which the complete

fish tank effluent is pumped first to the trickling filter and

sedimentation takes place only thereafter. This results in

maximum loads of organic matter (either expressed as

BOD or as COD) and SS, thereby leading to unnecessary

high BOD/Kjeldalh-N ratios, which may result in

reduced nitrogen removal capacity.

The relation for filter efficiency was corrected for

temperature (Box 1, Step 5) to extend the temperature

range at which the equation can be applied. The design

and Mayo (1972, 1974) and Wheaton (1977).

the water flow rate that is reused is assumed to be

r O2 requirement fish culture tank and TAN control.

h tank (Climit,TAN). When oxygen flow is chosen for

to be calculated for this flow.

) due to recirculation:

Þ

filter influent concentration when a filter is used as

ow rate (m3/day)).

ulation factor; R: recycle percentage (as decimal).

day).

monia removal of E at a certain temperature:

temperature (8C)

e ðhÞÞ�

day

24 h

��1

media void volume ðfractionÞ

�

edia ðm2=m3ÞÞay.

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 253

Fig. 4. Relation between the choice of k (Table 5), system volume, flow rate and biofilter surface area (Table 5) for a feeding period of 12 h

(European eel), PTAN = 47 g/kg feed, Climit,TAN = 6 g TAN/m3, C�TAN ¼ 2:0g=m3 and pH 7.5, rmax = 0.55 g TAN/m2/day and a HSL of 150–200 m3/

m2/day (after Heinsbroek and Kamstra (1990)).

procedure (Box 1) is described in detail by Wheaton

(1977, 1997).

4.5. Explanatory relationships

Another way to determine the flow rate and biofilter

dimensions for TAN control is presented by Bovendeur

et al. (1987) and Heinsbroek and Kamstra (1990).

Bovendeur et al. (1987) developed a design concept for

water recirculation systems, based on the dynamics of

waste production coupled to the 1/2-order and 0-order

kinetics of TAN removal (Harremoes, 1978). The

starting point for this design philosophy is the waste

production (g/day) and its diurnal variation. Together

with the water quality limits of the fish (Climit), these

factors determine the required flow through the fish

tanks. The design of the suspended solids removal unit

in turn is based on the maximum flow. The flow through

the fish tanks determines the concentration of waste

products and the amplitude of their diurnal variations.

This in turn determines the type of performance of the

biological reactor and thus the required specific surface

area to be installed in this filter.

In this concept, flow rates for oxygen, CO2 and total

solids are calculated using Eq. (4.11) for k = dmax and

PAcc = 0 (see Section 2.2). This further enables to

calculate the flow rate needed to control the TAN

Table 5

The relation between k, the fraction of PTAN (g/day) accumulating in the syste

rmax = 0.55 g TAN removal per m2/day and dmax = 2.1

k FractionRðd � kÞ dt A (m2)

1 0.25 85

1.3 0.15 111

1.5 0.08 128

Data are based on a 12 h feeding period for eel (from Heinsbroek and Kam

concentration and biofilter size using Eq. (4.12) and

Eq. (4.13):

Qr ¼���� k � P�

Climit � Cin � PAcc

Vsystem

� ����ðm3=dayÞ; (4.11)

Qr;TAN ¼k � PTAN�

Climit;TAN � C�TAN �PAcc;TAN

Vsystem

� ðm3=dayÞ;

PAcc;TAN ¼ PTAN

Z t2

t1

ðd � kÞ dt (4.12)

Abiofilter ¼PTAN�

rTAN

k

� ðm2Þ (4.13)

The design value k in both equations is used to tune

the dynamics in waste production to the TAN removal

kinetics in the biofilter and determines the intended

trickling filter performance. The choice of the design

value k (see Section 2.2) determines the fraction of the

TAN production temporarily accumulating in the

system volume and the trickling filter surface area to

m volume and biofilter surface area (A), for P = 47 g TAN/kg feed/day,

k FractionRðd � kÞ dt A (m2)

1.8 0.03 154

2.0 0.01 171

2.1 0 179

stra, 1990).

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260254

Box 2. Plug-flow model Nijhof (1995)

rTAN ¼ affiffiffiffiffiffiffiffiffiffiffiffi½TAN

p � b (4.14)

be installed (Table 5). Based on the choice of k in

Eqs. (4.12) and (4.13) three trickling filter performances

can be distinguished (Table 5):

a ¼ 7:81� 10�4HSLþ 0:2 (4.15)

1. A

b ¼ 0:1 (4.16)

rTAN ¼ c (4.17)

trickling filter design according to 0-order TAN

removal rate kinetics in relation to CTAN

(CTANC�TAN; k ¼ 1 and r ¼ r�max). This results in

a relatively small filter (Table 5), a constant high

TAN removal rate and a relatively large system

volume (Fig. 4). The system is vulnerable to TAN

peak loads and does not incorporate a safety factor.

½TAN� ¼ 5� 1:25h (4.18)

2. A

TANremoval � TANload ¼Q½TAN0

Ahðg=m2=dayÞ

(4.19)

d½TANh ¼ rTAN � A (4.20)

trickling filter designed according to 1/2-order

TAN removal rate kinetics in relation to CTAN as a

result of TAN diffusion limitation (CTAN <C�TAN;k ¼ dmax and r ¼ r�max) results in a large filter

(Table 4), a small system volume (Fig. 4) a low

average TAN removal rate but high stability and a

high safety factor. The flow rate is calculated using

Qr;TAN ¼ dmax � PTAN=C�TAN.

dh Q

3. A

½TANh ¼Z hþdh

h

d½TANdh

Dh ¼Z hþdh

h

rTANA

QDh

(4.21)

rTAN: TAN removal rate (g/m2/day); a: 1/2-order

coefficient (m/day); b: intercept (g/m2/day); a and

b are values depending on external factors (e.g.,

temperature, salinity, pH, [O2]) or internal prop-

erties (e.g. biofilm thickness, abundance of nitri-

fying bacteria, adaptation to specific

circumstances); c: 0-order removal rate

[TAN] > [TAN]*, the value c is determined by a

and b and the oxygen concentration; [TAN]h:

total ammonia concentration at depth h (g/m3);

[TAN]*: TAN concentration at which 1/2-order

TAN nitrification kinetics transfers into 0-order

kinetics or vice versa; h: height in filter bed (m),

A: biofilm surface area per unit h (m2/m); Q:

water flow through the biofilter (m3/day). Pilot

scale trickling filter characteristics; filter media:

Filterpak CR50, Mass Transfer Int., Heversham,

Cumbria, UK; specific surface area: 200 m2/m3;

void fraction: 0.93; random filter media; filter

height: 2.5 m and diameter: 1.2 m, ventilation

rate: 7000–7700 m3/day) (from Nijhof, 1995).

trickling filter design based on alternating 1/2-

order and 0-order TAN removal kinetics (alternating

CTANC�TAN and CTAN <C�TAN; 1 < k < dmax;

r ¼ r�max) combines the advantages of 1 and 2.

In well-ventilated trickling filters oxygen is

assumed not to be an limiting factor. The conditions

leading to each of the filter performances are presented

in Fig. 4.

4.6. A plug-flow model for nitrifying trickling filters

Nijhof (1995) described the nitrification perfor-

mance of a trickling filter in a plug-flow model using

plug-flow characteristics, TAN influent concentrations,

1/2-order/0-order model, hydraulic surface loads (m3/

m2 cross-sectional filter area/day) and the observed

stratification in TAN oxidation capacity (Box 2) which

was also observed by Boller and Gudjer (1986).

Model development was based on biofilm monitor-

ing studies, using biofilm samples from equidistant

places across the height of the trickling filter. For the

development of this model the effect of three hydraulic

surface loads (HSL = 75, 150 and 300 m3/m2/day) on

the 1/2-order coefficient a and intercept b (g/m2/day)

was determined. Nijhof (1995) found a = 7.81 � 10�4

HSL + 0.261 and b = 0.1 g/m2/day, thus giving fair

predictions of the trickling filter performance in the

pilot recirculation system. A ‘perfect’ description of the

trickling filter could be obtained by replacing the value

0.261 by 0.2. The better description was explained by

the scaling effect: the media in the pilot scale trickling

filter generally had a lower wetting when compared

with the wetting of biofilm samples of this filter when

used for parameterization of the model.

The observed positive relationship between the 1/2-

order coefficient a and the HSL was explained by an

increased wetting of the filter media at higher HSLs

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 255

because, at higher hydraulic loads, more water is

retained in the trickling filter. Hydraulic biofilm loading

rate (m3/m2 installed biofilter surface area) in combina-

tion with TAN concentration proved to be a key

parameter for predicting TAN removal, and thus, for

trickling filter design. The model (Eqs. (4.14)–(4.21))

showed that nitrification was only moderately affected

by shape and size of the trickling filter when the flow

rate remained the same.

However, there are restrictions on the application of

the model: (1) Eq. (4.14) predicts an unrealistically high

removal rate when higher hydraulic surface loads are

used. For the hydraulic surface loads of 469 and

1231 m3/m2/day, Greiner and Timmons (1998) could

not detect any effect on nitrification; (2) Eq. (4.18)

predicts no TAN removal for a filter higher than 4 m

(Nijhof, 1995).

Based on his results, Nijhof (1995) concluded that

the concept of Bovendeur et al. (1987) overestimated

the TAN removal rate, especially at the lower hydraulic

surface load of 75 m3/m2/day and underestimated the

removal rate at a HSL of 300 m3/m2/day, due to the

assumption of a completely mixed reactor. Mixed

conditions can probably be only assumed for small

filters at high hydraulic surface loads.

Kamstra et al. (1998) validated the model of Nijhof

(1995) after prior adjustment of model parameters in 70

full scale trickling filters in 14 commercial eel farms.

They did not include the effect of stratification in TAN

oxidation capacity in their calculations (Eq. (4.18)). All

filters in their research were operating at filter effluent

concentrations �1.8 g TAN/m3, indicating that none of

the trickling filters were likely to be operating at 0-order

removal kinetics only (CTAN <C�TAN) but, instead, on 1/

2-order and 0-order removal kinetics or 1/2-order TAN

removal kinetics.

For the observed TAN removal rates in commercial

filters, the removal rate predicted by the model needed

to be adjusted for the type of biofilter media applied:

- r

observed (g TAN/m2/day) = �0.03 + 1.27rpredicted

(r2 = 0.72; cross flow media, Munters);

- r

observed (g TAN/m2/day) = �0.01 + 0.69rpredicted

(r2 = 0.69; vertical flow media, Bionet);

- r

observed (g TAN/m2/day) = 0.03 + 0.54rpredicted

(r2 = 0.68; random flow media, Filterpac CR50).

Although the model outcomes had to be corrected for

full-scale situations, they observed that the model

sufficiently explained the TAN removal rate for a wide

range of trickling filters. When compared with vertical

flow (Bionet) and random flow (Filterpac media), farms

applying cross flow media (Munters) showed the best

TAN removal rates comparable to what is reported in

literature (Kamstra et al., 1998).

However, the high removal rates reported for cross

flow media were also related to differences in system

configuration and operation:

- s

uspended solids were removed by a submerged filter

instead of screen filtration which might have lowered

the BOD loading rate on the cross flow media;

- h

igh hydraulic surface loads were applied for this

media which might have had a positive effect on

preventing clogging as well as on biofilm character-

istics;

- a

ir in the trickling filter was exchanged by forced

ventilation instead of natural ventilation and might

have somewhat improved the oxygen transfer.

For the three applied filter media types in

commercial farms and a range of hydraulic surface

loading conditions, the observed TAN removal rate

could also be described as a function of the TAN loading

rate (L): robserved (g TAN/m2/day) = 0.01 + 0.32L(r2 = 0.798). It shows that the TAN loading rate is an

important parameter for predicting TAN removal rates

in trickling filters (Kamstra et al., 1998). The highest

observed TAN removal rate for a trickling filter was

1.1 g TAN/m2/day.

Kamstra et al. (1998) tested the power of the model

by predicting the diurnal and daily variations in TAN

concentrations observed in commercial farms. The

prediction of TAN removal was only satisfactory for

two out of four farms. This result was explained by the

use of demand feeders in those farms where predictions

were incorrect. The use of demand feeders made it

difficult to predict the instantaneous feed consumption

and thus also the TAN production. However, the

utilization of the feed may have also fluctuated over

time making it difficult to estimate the exact TAN

production. The model equations can be applied in a

spreadsheet program and allows one to predict the

effluent TAN concentration, the average and maximum

filter TAN removal rate, the treatment efficiency and the

conditions giving maximum removal rate.

The model was also used to predict the energy costs

for pumping water across the filter, for a range of flow

rates, feed loads and allowed filter influent TAN

concentrations (Kamstra et al., 1998). Although many

other variables may influence the nitrification rate (O2,

COD load, pH), the authors believed that incorporation

of these variables in the model would only result in

small model improvements.

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260256

4.7. Some practical aspects of trickling filter designand operation

The design procedure outlined earlier, matches the

output of ammonia by fish with a certain volume of

trickling filter material, taking into account the

ammonia removal rate and the specific surface area

of the filter material used. The exact dimensions of the

filter in terms of height and surface are still to be

determined.

In practice, the height of the filter bed in trickling

filters can vary between 0.6 and 4.5 m (Kamstra et al.,

1998). In order to prevent clogging, a hydraulic surface

load in the order of 300–400 m/day is required. This

implies that relatively large flows need to be generated

in shallow filters. Therefore, it is advisable to avoid

shallow filters and to target filter heights of 2–4 m. To

prevent deformation of the stacked plastic media in the

lower part of the trickling filter, different grades

(thickness) of filter media can be used or supports

should be installed at distinct depths.

High void ratios reduce clogging. A definition of

void ratio is given by Wheaton et al. (1994) as ‘‘the

volume of air left in a filter after it is filled with media

divided by the total volume of the empty filter up to the

same level as the media will fill it’’. Boller and Gudjer

(1986) judged a specific surface area of 150–200 m2/m3

to be most suitable for the corrugated plastic media they

applied in their wastewater treatment research. Similar

specific surface areas for plastic media (Bionet 160 m2/

m3; Filterpac 200 m2/m3 and Munters 234 m2/m3) are

installed in trickling filters applied in aquaculture

(Kamstra et al., 1998).

The trickling filter should allow space at the top for a

water distribution device and should be open at the

bottom to assure optimal ventilation. In some designs,

the trickling filter is also used as a header tank for

Fig. 5. Three types of filter media applied in trickling filters (from left to righ

m3, void fraction 0.93, diameter �5 cm, height �3.2 cm, Mass Transf

0.50 m � 0.50 m � 0.60 m (length � depth � height), specific surface area 2

and cross flow media (FKP319, 0.30 m � 0.30 m � 2.40m, specific surface

Netherlands).

further distribution of water to the fish tanks and is

closed at the bottom. In these designs, a blower has to be

installed for forced ventilation.

Apart from nitrification and removal of BOD, a

trickling filter is ideally suited for removal of carbon

dioxide. Moreover, it can be used for evaporation

cooling in warm climates. In both cases, a controlled

airflow over the filter is needed. To realise this, the space

on top of the trickling filter can be closed and connected

to a ventilation system. For optimal degassing, the

minimum ratio of air to water flow needed is in the order

of 10, while a minimum filter bed height is needed.

When higher ventilation rates are applied, the increased

evaporation may help in cooling the water during

summer time. Forced ventilation also helps in prevent-

ing stagnant air in periods when the water temperature

in the filter is almost similar to the air temperature

outside the filter. Stagnant air reduces the oxygen partial

pressure and results in poor aeration of the bulk water,

which may subsequently reduce the nitrification

capacity of the filter (Wik, 2003). The type of filter

medium has an effect on the specific removal rate of

ammonia (Kamstra et al., 1998). Cross flow media

perform better than vertical flow or random flow

media—an effect which is attributed to differences in

hydraulic and wetting characteristics (Fig. 5). Clogging

of filter media can be a serious problem in commercial

farms and must be avoided. In this respect, the effect of

the hydraulic surface load of the filter and the type of

filter material are difficult to quantify. Experience has

shown that random flow media are prone to clogging,

which is the reason why vertical flow and cross flow

media have become more popular.

Cross flow and vertical flow media come as self-

supporting blocks, which can be stacked easily and taken

out when necessary. Random media are mostly in the

form of loose ‘balls’ and require a special support frame.

t): random flow media (Filterpak1-CR50: specific surface area 200 m2/

er Int., Heversham Cumbria, UK); vertical flow media (Bionet1,

00 m2/m3, void fraction 0.95, Catvis BV, Den Bosch, The Netherlands)

area 150 m2/m3, void fraction 0.92, Fleuren & Nooijen, Someren, The

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260 257

A good water distribution device on top of the filter

is essential to utilise the total filter volume. Water can

be distributed through a moving arm, a perforated

screen or a nozzle. In round filters with random media,

a rotating beam is often applied. These constructions

are sensitive to mechanical wear and need to be

constructed carefully. Perforated screens are often

used on small filters, but require frequent maintenance

to avoid clogging of the holes. Nozzles (rotating) can

handle large flows (Summerfelt et al., 2001) at little

head pressure and can provide effective water

distribution.

In recirculation systems without denitrification, the

pH drop caused by nitrification has to be counteracted.

Although most fish species appear to be tolerant to

relatively low pH levels, the nitrification rate is

substantially reduced at values below pH 7 shown by

elevated TAN and nitrite concentrations at low pH.

Sodium bicarbonate is usually added to stabilise the

pH at values around 7. In the Netherlands, many

European eel and African catfish farms operate their

systems at low pH and high TAN and nitrite levels.

Heinsbroek and Kamstra (1995) observed pH values of

6–9; TAN concentrations of 0–40 g TAN/m3 and

nitrite concentrations of 0–20 g NO2-N/m3 for com-

mercial eel farms.

Trickling filters are robust and can be easily taken

off-line for a few hours without problems. In submerged

filters, the risk of anoxia through pump failure and

subsequent damage to the biofilm should be taken into

consideration. Whatever the filter mode, caution needs

to be exercized when fish are treated by a bath treatment

against diseases.

In general, a trickling filter performs optimally at

increasing or stable waste loads (up to the designed

maximum load). When the feed load in the system is

reduced (e.g. by harvesting fish), part of the filter may

be detached. The filter sheds part of its biofilm and the

detached biofilm particles add to the SS concentration in

the water. Since they generally have a size below

40 mm, they are difficult to remove. This results in a

strong increase in fine suspended solids, which can

hamper fish performance.

The process of biofilter detachment is not yet fully

understood. Biofilm parameters, which were used to

clarify the process of biofilm detachment, are dry

density, wet density, the content of the extra cellular

biopolymer (ECP), increased gas content in maturing

biofilm and shear stress (Ohashi and Harada, 1994).

Further research is required to improve current knowl-

edge on managing biofilm thickness and biofilm

detachment in trickling filters.

5. Conclusions

In evaluating the presented material in biofilter

design and operation, it can be concluded that:

(1) d

etermination of the waste production is very simple

when applying a mass balance analysis at the level

of the fish organism, showing the utilization of

dietary feed by fish (Table 1). However, only in a

few studies is this approach integrated into the

design procedure;

(2) th

e water quality demands are a weak point in the

whole design cycle. This concerns two aspects: the

chronic effects of average daily values and the effects

of fluctuation in water quality during a 24-h cycle;

(3) im

provement of TAN removal rate (g TAN/m2/day)

of trickling filters seems to be possible when

structurally low C/N ratios can be obtained

(Bovendeur, 1989; Zhu and Chen, 2001). The

higher observed TAN removal rates at low COD/N

ratios in tertiary nitrifying filters applied in waste-

water treatment confirm this capability (Gujer and

Boller, 1986; Lazarova et al., 1997; Metcalf and

Eddy Inc., 2003). The potential reduction of C/N

ratios in fish waste (metabolites) can be supported

by diets incorporating highly digestible nutrients –

formulated at optimal protein/energy ratios – with a

low dust fraction and which stimulate the produc-

tion of fecal pellets with a high consistency.

Selection and integration of a more efficient solids

removal process may lower C/N ratios further.

Improvement of biofilter performance will, there-

fore, require the input of system engineers and

nutritionists (Piedrahita, 2003);

(4) b

ased on laboratory- and pilot-scale studies, a

number of models are available for the design of

trickling filters. However, model validation on a

commercial scale is lacking. Evaluation of com-

mercial scale recirculation systems in relation to the

applied design concepts would help to improve our

knowledge of filter design, water quality control and

applied safety factors;

(5) m

any commercial trickling filter designs are based

on the empirical experience of a few companies

with a given recirculation system configuration for

the production of a specific fish species. Since many

factors affect trickling filter design (e.g. fish species,

feed composition, feeding strategy, system config-

uration, biofilter configuration, type of media, feed

management, etc.), changing these factors should

be first tested on a pilot scale before applying it

commercially;

E.H. Eding et al. / Aquacultural Engineering 34 (2006) 234–260258

(6) th

e operation and management of trickling filters

mainly focuses on the prevention of clogging and

the optimization of biofilm stability.

Acknowledgment

The authors would like to thank Oliver Schneider for

critical reading of the manuscript.

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