4 diseño y operación de filtro percolador
DESCRIPTION
ARTICULOTRANSCRIPT
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: [email protected] (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 and4.34 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-Noxidized 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 and8.33 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 (TANdiffusion 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 removalkinetics in relation to CTAN due to diurnal variation
in TAN concentration;
4. 1
/2-order removal kinetics in relation to CO2(oxygendiffusion 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 DO2are the corresponding
diffusion coefficients (at 25 8C and pure water, see
Gujer and Boller, 1986) and
� v
O2;TAN is the stoichiometric constant for oxygenconsumption 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=
33
–3
4%
,te
mp
erat
ure
=2
48C
,
BO
D=
25
g/m
3,
TA
N=
5g
/m3
Tri
ckli
ng
filt
erm
od
e;
sim
ilar
con
dit
ion
sas
inth
ead
apta
tio
np
erio
dex
cep
t;
full
-gro
wn
bio
film
,C
TA
N=
0–
7g
/m3
Sea
wat
er(p
H8
.3;
O2
8.4
g/m
3;
248C
)r T
AN
=0
.23ffiffiffiffi Cp�
0:1
11
/2-o
rder
TA
N
Sea
wat
er(p
H8
.3;
O2
8.4
g/m
3;
248C
)r T
AN
=0
.28;
C� T
AN¼
3g=m
30
-ord
erO
xy
gen
Fre
shw
ater
(pH
8.2
;O
28
.4g
/m3;
248C
)r T
AN
=0
.55ffiffiffiffi Cp�
0:1
21
/2-o
rder
TA
N
Fre
shw
ater
(pH
8.2
;O
28
.4g
/m3;
248C
)r T
AN
=0
.69;
C� T
AN¼
2:2
g=m
30
-ord
erO
xy
gen
Sta
nd
ard
bio
film
adap
tati
on
con
dit
ion
sw
ere:
tric
kli
ng
filt
erm
od
e,b
iofi
lmsp
ecifi
csu
rfac
ear
ea1
.1–0
.82
m2;
HS
L=
15
0–2
00
m3/m
2/d
ay;
Tem
per
atu
re=
248C
;T
AN
load
ing
rate
circ
a0
.5g
TA
N/
m2/d
ay;
DO
=7
g/m
3;
pH
=7
;C
OD
load
circ
a5
gC
OD
/m2/d
ay(b
atch
wis
efe
dw
ith
pre
-dig
este
dfi
shfe
ed),
BO
D/C
OD
rati
o=
0.8
,b
icar
bon
ate
was
add
edfo
rp
Hco
ntr
ol.
(a)
Aver
age
TA
Nre
moval
val
ues
for
bio
film
sam
ple
sof
atr
outa
nd
catfi
shpil
ots
cale
reci
rcula
tion
syst
em(B
oven
deu
ret
al.,
19
87,B
oven
deu
ran
dK
lap
wij
k,1
98
6).
(b)
Long-t
erm
effe
ctof
accu
mula
ting
bio
film
subst
rate
due
to
CO
Dlo
adin
gra
tean
def
fect
of
pH
on
TA
Nan
dN
O2-N
rem
oval
rate
(Boven
deu
r,1
98
9).
(c)
Sh
ort
-ter
mef
fect
of
CO
Dlo
adin
gra
te(B
oven
deu
ret
al.,
19
90).
(d)
Eff
ecto
fsa
lin
ity
on
TA
Nre
moval
rate
(Nij
ho
fan
dB
oven
deu
r,1
99
0).
aA
pp
aren
t0
-ord
erd
ue
toth
ep
rese
nce
of
amm
onia
.b
0-o
rder
NO
2-N
rem
oval
rate
po
ssib
lyd
ue
toin
hib
itio
nb
yu
nio
niz
edn
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 bydiurnal variation in waste production due to the
feeding strategies applied;
� s
tructural high organic matter loading rates due to, forexample, 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 fractionof 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 thehydraulic requirements for trickling filter operation.
� T
AN concentrations in the bulk water higher thanC�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 waterquality 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. Ab 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. ATANremoval � 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 Q3. 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 filterinstead of screen filtration which might have lowered
the BOD loading rate on the cross flow media;
- h
igh hydraulic surface loads were applied for thismedia 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 forcedventilation 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 simplewhen 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 thewhole 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, anumber 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 basedon 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 filtersmainly 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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