4. Methods for treatment of pulp and
paper production effluentsBiological treatment, typical applying the activated sludge system, is the standard method for the purification of pulp and paper mill effluents. In case of severe effluent limits, regarding organic pollution, additional chemical treatment may become necessary.
For new units planning the production and developing a treatment concept takes place at the same time so that in modern pulp production complexes the philosophy of "preventing before treating" is mostly realised.
4.1 Analysis of the different effluent
streams
The multi-stage process from wood to paper produces several effluent streams
with different grades of pollution (see chapter 3.1 production process). This
paper deals with lab-scale investigations to improve the "end of pipe"
treatment in Sari. Therefore, analysing the different wastewater streams is
inevitable to reach the primary objective.
In order to select a both economical and efficient treatment, sequence assessment
of the individual substreams is necessary.
It has been shown that more than half of the lignin derived pollutants in
pulp mill effluents and resin acids are likely to be non-biodegradable or
even harmful to aquatic life. (Römpp, 1992).
4.1.1
Identification
According to the different production stages, the wastewater flow can be divided
up into the following main streams (see Figure 3 and Figure 4).
Figure 3 - Main-streams
Figure 4 - Critical Sub-streams
The water from the Wood Room, Plug Screw Feeder, Chip Washing, CMP Bleaching
and the pulp Machine 2 (PM2) effluent emerges during the pulp production,
whereas the wastewater streams PM1 and PM2 originate from the paper production.
The Wood Room Effluent with a COD load of 280 kg/d can be considered as a
relatively uncritical sub-stream. Furthermore, the COD concentration in the
effluent of PM1 is quite low (2,157 mg/l). The Bleaching wastewater stream
can be neglected as well because of its low flow.
Hence, the COD load released by the pulp mill averages 27 % of the total load.
The highest COD concentrations are to and in the effluents of the Plug Screw
Feeder and Chip Washing (concentration range: 11,000 to 20,000 mg COD/l).
The biggest part of the total COD load, about 21,500 kg COD/d, is due to the
PM2 effluents.
Separate pre-treatment of the Plug Screw Feeder, Chip Washing and the PM2
effluents appeared to be the most economised approach to achieve the overall
COD effluent limit of 100 mg/l.
4.1.2
Assessment of treatment efficiencies
In order to assess the impact of separate pre-treatment of the three selected
substreams (Plug Screw Feeder, Chip Washing and the PM2 effluents) on the
overall efficiency of wastewater treatment, balancing calculations were carried
out.
Based on previous measurements of the Canadian company in Sari, a comparison of
the COD load with and without separate treatment should give a first indication
which percentage of the sub-stream COD has to be eliminated in order to reach
the requested limits. As available data from the site show major flow and
load vary with the time, the assessment had to be based on reasonable assumptions
(see Figure 5).
As the production process itself or the sort of wood utilised has a big impact on the effluent composition, theses factors must be taken into consideration as well. Generally, using softwood, for instance poplar, results in lower COD loads. With hardwood like oak the COD concentration almost doubles.
Figure 5 - Comparison of PM2 effluents
According to the performance data available from the Sari treatment plant
typically over 98 % of the BOD load is eliminated but only 87 % of the COD
load.
In Figure 6 a "simulation" of sub-stream treatment is outlined based
on different measurements.
A subtraction of the sub-stream COD loads (A, B) from the mainstream (C, D)
and an assumed treatment of 87 % to 92 % of the residual stream give a potential
goal for the sub-stream treatment (A*, B*). The "simulation" shows
that it will be very difficult to reach the requested value by separating
only these sub-streams. In the sub-streams the COD elimination would have
to reach a value between 98 % (E) and 97 % (F). But it should not be a problem
to add an additional stream as a part of the pulp mill effluent if the authority
asks for even stricter effluent values.
In the selected sub-streams the portion of non-biodegradable substances was higher than in the other wastewater streams, as indicated by the respective BOD/COD ratios.
Figure 6 - Assessmentof treatment efficiencies
4.1.3 Chemical
properties of the pulping effluent
The main compounds in the sulfite pulping effluents are the ligninsulfonates
as listed below in Figure 7. The specific COD load from hard wood is 2.5 times
higher than the one from softwood. Reason for this is the different composition
of hemicellulose and lignin.
Figure 7 - Main compounds
Generally, the main compounds in pulp and paper mill effluents are
o ligninsulfonates which are non-biodegradable,
o organic acids and saccharine which can be degraded in an aerobic process.
The single compounds of pulp and paper effluents were analysed via ultrafiltration
and aluminium oxide (Al2O3) adsorption by the Papiertechnische Stiftung (PTS).
Figure 8 points out the volume distribution of the main ingredients. In Figure
9 the molecular weight distribution is shown. From this it can be seen that
the ligninsulfonates and lignin itself have a molecular size of 0.01 - 0.1
µm and a molecular weight of 20,000 - 100,000 D on average. In general,
the total amount of these compounds equals 30 % (Demel, 2000).
A further identification showed that about 50 % of the total content belongs
to the group of lignin and humic acids with a molecular weight less than 5000
D and a size of 0.01 µm, respectively.
The presence of these substances can be considered as the main reason for
the poor biodegradability of pulp and paper mill effluents.
Figure 8 - Molecular size distribution
Figure 9 - Molecular weight distribution
The word lignin is derived from the Latin word lignum meaning wood. Lignin
is an natural polymer made up of phenylpropane units (Figure 10). Lignin performs
multiple functions that are essential to the life of the plant. By decreasing
the permeation of water across the cell walls in the conduction xylem tissues,
lignin plays an important role in the internal transport of water, nutrients
and metabolites. It imparts rigidity to the cell walls and acts as binder
between the wood ingredients. It is creating a composition material that is
strongly resistant to impact, compression and bending. Lignified tissues effectively
resist attack by microorganisms. Depending on the sort of wood lignin has
a different composition. Basically lignin is formed in a woody plant by dehydrogen
polymerization of three phenylpropanoid monomers
Figure 10 - Three lignin compounds
Figure 13 shows the structural model of lignin with the polymeric strings
hinted through the raster.
The physical properties like the molecular mass are described above. In wood
lignin behaves as an insoluble, three-dimensional network. Lignin exhibits
a glass transition temperature of about 160 °C.
Figure 13 - Structural model of lignin
The chemical properties depending on the pulping process are:
o Alkaline Pulping (Kraft Pulping)
o Sulfite Pulping
The Sari pulp production uses the Sulfite Pulping process for delignification.
Under Sulfite Pulping conditions, lignin is sulfonated and rendered water-soluble.
A major sulfonation reaction is demonstrated in Figure 11.
Figure 11 - Sulfonation reaction
Ligninsulfonates are heterogeneous in terms of their molecular polydispersity
and structure. They are soluble in water over the entire pH range, but insoluble
in ethanol, acetone and other common organic solvents.
For everybody visible is the yellowing reaction of lignin. It discolours in
the presence of alkali and oxygen (air), or when exposed to daylight, causing
the yellowing of paper or wood.
Microbial and enzymatic degradation of lignin is a complex biochemical process.
So to say lignin decays by "white-rot fungi".
In the lignin-degrading bleaching of pulp, bleaching chemicals like ozone,
hypochlorite and oxygen causes degradation of lignin polymer. This is exemplified
by the reaction of chlorine dioxide in Figure 12 (Römpp, 1992).
Figure 12 - Bleaching reaction of chlorine dioxide
The knowledge of the chemical, physical and biological properties of the pulp
and paper mill effluent ingredients is the general basis of further investigations
4.2 Treatment Concepts
After the analysis of the different wastewater streams and their compounds
a new treatment concept had to be developed. As described in the introduction
(Chapter 2) Sari is already equipped with a biological treatment stage.
The existing treatment plant reaches a BOD-degradation of up to 98 %, but
the reduction of the COD load of 87 % has to be improved. The high BOD elimination
shows that the existing plant successfully eliminates the biodegradable part
of the organic pollution load. However, refractory COD, caused by the lignin
derived species, as explained above, and partially contained in the selected
sub-streams, limits the COD removal achievable with this plant.
Figure 15 points out the present treatment concept based on an activated
sludge process.
A new treatment concept is developed based on the results of respective test
results.
In order to conduct the pre-studies 700 l of wastewater (selected sub-streams)
were air freighted from Sari to the Linde laboratory in Munich.
A principal scheme of all scheduled lab-scale tests is shown in Figure 14.
Figure 14 - Scheduled lab-scale tests
Figure 15 - Present treatment concept - new treatment concept
4.2.1 Biological
treatment
Today, the most common wastewater treatment procedure is biological, whereby
concentrated masses of microorganisms break down organic matter, resulting
in the stabilisation of organic wastes. These microorganisms are broadly classified
as aerobic, anaerobic or aerobic and anaerobic (facultative). The facultative
organisms may function in either an aerobic or anaerobic environment. The
aerobic organisms require molecular oxygen for metabolism; anaerobes derive
energy from organic compounds and function in absence of oxygen.
The predominant species used in biological systems are known as heterotrophic
microorganisms. These require an organic carbon source for both energy and
cell synthesis. Autotrophic organisms in contrast, use an inorganic carbon
source such as carbon dioxide or carbonate and derive energy from the oxidation
of inorganic compounds such as nitrogen, sulfur or from photosynthesis (Fox,
1976).
For the planned biological treatment two different approaches were tested. A pure oxygen activated sludge system, a subdivided part of the existing treatment plant and/or an external anaerobic reactor.
Pure
Oxygen Activated Sludge System
The activated sludge process is a continuous system in which microorganisms
grow under aerobic conditions and form so called activated sludge flocks.
These flocks are mixed with wastewater in the reactor and then physically
separated by gravity clarification.
The aerobic degradation of organic substances takes place almost exclusively
through bacteria and lower fungi. By using synthetic aeration and a high concentration
of microorganisms the degradation processes faster. The bacteria take up the
solute through their cellular membrane. With the presence of nutrients like
phosphorous and nitrogen the microorganism can produce proteins. With the
oxidation (combustion) of carbonated substances the cells can meet the energy
demand for the metabolism. Overall the final products of this reaction are
CO2, H2O and biomass (Möbius, 1997).
The concentrated sludge is recycled to the reactor and mixed with the incoming
waste. Oxygen is provided by oxygen-enriched air through surface aeration.
Figure 16 - Activated sludge system
Aerobic laboratory test plant
Even though measurements existed concerning the aerobic biological degradability
of the total effluents it was necessary to investigate the biodegradability
of every selected sub-stream. For that reason a small test plant was installed
at the Linde laboratory. The experimental setup is shown in Figure 19.
Figure 19 - Experimental setup - aerobic treatment
Figure 89 - photo - aerobivc experimental setup
The reactor consisted of an aeration tank with a reaction volume of about
5 l and an additional clarifier tank. Substrate was pumped from a storage
tank into the aeration basin. Compressed air or pure oxygen was utilized to
satisfy the oxygen demand of the microorganisms. Sensors monitored the oxygen
content in the reactor and the pH value, respectively. A controlled volume
of settled sludge in the clarifier was recycled into the aeration basin.
Start-up conditions
To save time and to get a working biological culture 10 days prior to the
arrival of the samples from Sari the test plant was started up with wastewater
and sludge from a pure oxygen operated activated sludge plant treating sulfite
pulp mill effluent at Mannheim Germany.
In the start up phase the flow rate was increased continuously up to 4-5 l/d.
The recycle rate was set to 5 l/d (100 %). Sometimes it was necessary to dose
some phosphoric acid in order to control the pH value as well as nutrient
addition. The oxygen supply was regulated between 3.6 and 4.5 mg/l.
Before getting realistic results about the treatment capacity the microorganisms
have to adapt to the given conditions. To avoid shock loadings it was decided
to gradually substitute the effluent from the German plant by the Sari wastewater.
Figure 20 shows the different influent concentrations.
The hydraulic retention time was set from 0.4 to 1.0 d in average due to the
higher influent concentrations and organic load, respectively. The analysis
of the influent water showed that it was necessary to dose extra nutrients
(ammonium phosphate, see Figure 18) to get optimal conditions for bacterial
growth.
Figure 18 - Nutrient addition
Figure 20 - Comparison of the influent parameters Mannheim - Sari
Wastewater composition
The wastewater samples from Sari included 400 l of PM2 effluents, 150 l of
Plug Screw Feeder clear filtrate and 150 l of Chip Washing effluent. Subsequent
chemical analysis verified the respective data submitted by the Canadian company
as listed below in Figure 17.
Figure 17 - Comparison of the analysis results
It is obvious that the BOD/COD ratio 0.23 of the PM2 effluent is the lowest
and the PM2 flow of 3,800 m³/d is the highest. For this reason the focus
was set on the investigations of this sub-stream. For the aerobic treatment
it was planned to use the PM2 effluent first and in a second step to proportionally
compose a wastewater mix from the three sub-streams according to the "real"
conditions.
Results
Six days after changing to PM2 effluent, a good adaptation of the biomass
was obtained. The COD concentration was reduced from 7,200 mg COD /l to 3,200
mg COD /l. A COD degradation of only 56 % is a typical value for poorly biodegradable
pulp and paper mill wastewater (see Figure 21).
Another typical effect for this kind of wastewater was observed after aerobic treatment. The colour of the water changed from beige to deep brown/black due to a partial oxidation of the polymeric lignin compounds.
During the aeration intervals intensive foaming was observed. As a consequence
the pure air aeration was replaced by an air and pure oxygen mixture.
The treatment parameters during the test period are listed in Figure 21.
Figure 21 - Aerobic treatment
The next step was to change the pilot influent from PM2 water only to a proportional
mix of the three wastewater streams. The influent COD concentration inclined
from 7,200 mg/l to 8,100 mg/l. Cause of the slight difference between the
influent COD values it was expected that the adaptation phase would complete
faster than before.
Finally similar to the pure PM2 water about 50 % of the COD could be reduced.
That was not surprising causes of the ingredients are more or less the same
like in the PM2 wastewater.
Anaerobic wastewater treatment
Anaerobic treatment is defined as biological oxidation of wastewater by microbes
in the absence of molecular oxygen. The activity of a complex mixture of microorganisms
involves degradation, transformation, and synthesis reactions of organic matter
and is finally leading to mineralization (Rintala, 1993).
Three basic groups of bacteria are involved in this multistage process. The
complex organic compounds are sequentially converted, through a series of
intermediate compounds, to methane and carbon dioxide, as indicated by the
four-step process shown in Figure 22 (Möbius, 1997).
Figure 22 - Four stage of anaerobic metabolism
Complex, high molecular weight, soluble organic compounds (carbohydrates,
proteins) must first be hydrolysed (Stage1) to simple organics (simple sugars,
amino acids, glycerol, fatty acids).
These simple organics are converted by acid-forming bacteria to higher organic
acids and to acetic acid, hydrogen and carbon dioxide in a fermentation or
acidogenic phase (Stage 2).
The higher organic acids are subsequently transformed to acetic acid and hydrogen
(Stage 3) by acetogentic bacteria. The acidogenic and acetogentic bacteria
belong to a large, diverse group that includes both facultative and strict
anaerobes. Depending on the wastewater characteristic one of the groups predominates.
The final step (Stage 4) to produce methane is carried out by three groups
of methane bacteria. These strict anaerobes are capable of metabolizing formic
acid, methanol and carbon monoxide, as well as acetic acid, hydrogen, and
carbon dioxide to methane (Lee, 1987).
It is known that in anaerobic processes where inorganic sulfur is a ingredient
of wastewater, the sulphate reducing bacteria are very important. Sulfate
and sulfite is present in the effluents from neutral sulfite semi-chemical
(NSSC) or chemi-mechanical (CMP) pulp mills (Umweltbundesamt, 1995).
The Sari pulp and paper production uses these pulping methods. The sulfur
reducing bacteria use sulfate and sulfite as electron acceptors in their metabolism.
Sulfur reduction can become a significant factor in the performance and operation
of pulp and paper anaerobic treatment. The hydrogen sulfide produced can be
both toxic and corrosive. Generally, the hydrogen sulfide dissociates in water
in two steps. The species presents depends on the pH, as indicated in Figure
23.
Figure 23 - Sulfide species as function of pH
Undissociated H2S is the most toxic sulfide species. Inhibition can be minimised
by increasing the pH (Lee, 1987).
It is obvious that the anaerobic metabolism is a sensitive system which has
complex pretension to process variables. On that account the pH should be
regulated between pH 9.7 and 7.4. Another important characteristic of this
process is the operating temperature of 32 - 37 °C (Fels, 1997).
Anaerobic laboratory Test Plant
Although the recent plant configuration in Sari does not include an anaerobic
treatment stage, it is known from literature that especially pulp and paper
mill effluents are suitable for anaerobic treatment (Kortekaas, 1998).
Hence, the Chip Washing effluent and the Plug Screw Feeder clear filtrate
were chosen because of their high COD concentrations of 11,000 - 20,000 mg
COD/l and low flow of 300 m³/d, respectively.
As a third step, a 1:1 mixture of both wastewater streams was composed as
influent. In order to avoid a breakdown of the sensitive biocenosis due to
the high acidity of this mixture it was necessary to neutralise the influent
prior to feeding.
The set-up of the anaerobic test plant is shown in Figure 25.
Figure 25 - Experimental setup - anaerobic treatment
Figure 90 - photo - anaerobic experimental setup
Start-up conditions
As described above an anaerobic system is very sensitive towards significant
milieu changes. For this reason a consequent pH and temperature control is
obligatory.
In general, the residence time in an anaerobic reactor is higher than in an
aerobic reactor due to the lower growth and conversion rates, respectively.
Thus, the flow was regulated to 1.5 l/d in average. To start up the anaerobic
treatment as fast as possible anaerobic sludge from the municipal treatment
plant in Starnberg (Munich) was utilised as an inocculum.
The general procedure was similar to the start-up of the aerobic treatment
tests (see Chapter 4.2.1 pure oxygen wastewater treatment). Subsequently,
the anaerobic reactor was fed with different wastewater compositions: a mixture
of PM2 and Mannheim wastewater in the beginning followed by pure PM2 effluents
after 10 days of stable operation - indication for a reasonable biomass adaptation.
Results
The COD reduction of 50 % under anaerobic conditions was not significantly
lower compared to the aerobic treatment (55 %). The COD concentration decreased
from 7,200 to 3,600 mg COD/l (see Figure 26).
Figure 26 - Anaerobic treatment
The gas production, a measure of the suitability of wastewater for anaerobic
treatment, amounted to 0.8 l/d. The composition of the gas indicates that
the wastewater did not contain serious inhibitor substances. The composition
of gas produced is presented in Figure 24.
Figure 24 - Gas composition
4.2.2 Chemical
treatment
Even though biological treatment of the selected substreams reduces the COD
in the range of 50 - 60 % only, it is an indispensable first treatment step,
as it limits the consumption of more costly utilities by chemical treatment.
The chemical treatment of pulp and paper mill effluent represents a traditional
and well-known method.
In Sari, a subsequent chemical treatment stage after the biological stage
was already installed, but the expected results could not be observed. Figure
27 shows the results of further investigations by the Canadian company
Figure 27 - Flocculation and Coagulation test with the total effluents
It is obvious that the amount of chemicals needed would be intolerable high.
On the one hand the strong dilution of the wastewater stream could have caused
a weakening of the coagulation effect; on the other hand it is possible that
a part of the effluent was unsuitable for this kind of chemical treatment.
For this reason the suitability of the highly concentrated sub-streams for
chemical treatment methods had to be investigated.
The practicability of different methods, for flocculation, adsorption and
oxidation was tested in lab-scale experiments.
Coagulation
and Flocculation
The stability of a suspension depends on the number size, size, density, surface
properties of the solid particles and on the density of the external phase
or dispersion medium.
In aqueous suspensions, the particle surface has an electrical, usually negative
charge. If counterions, for instance Ca2+, are present in the surrounding
water, they accumulate on the surface of the suspended particles, forming
an ionic double layer. The excess negative charge at the shear surface of
the double layer, the zeta potential, can be measured. As the zeta potential
increases, the coulombic repulsion between the particles becomes stronger
and the suspension is more stabile.
Four models are currently used to explain how flocculants aid particle agglomerate:
Double layer compression (Coagulation)
Specific ion adsorption (Coagulation)
Polymer charge patch (Flocculation)
Polymer bridging (Flocculation)
However, more than one of these mechanisms probably act simultaneously (Ullmanns,
1988).
The coagulation and flocculation are similar and their definitions vary.
Coagulation is defined as the agglomeration of suspended particles due to
the reduction of the repulsive forces caused by surface charge that keeps
them separate and suspended in liquid medium (Fox, 1976). The repulsive forces
are reduced either by addition of inorganic electrolytes, which shield the
repulsive surface charges, or by addition of polyelectrolytes that bind to
and neutralise the surface charge.
Flocculation is the agglomeration of particles due to the bridging effect
exerted by polymers that are adsorbed to more than one particle. Often both
mechanisms occur simultaneously when polymeric polyelectrolytes are involved.
Flocculated aggregates tend to be more porous than coagulated aggregates (Ullmanns,
1988).
Fe (III) Chloride, Aluminium Sulfate (Alum), Polyaluminiumchloride
(PAC)
Fe (III) Chloride or Alum is frequently used as coagulant because of the relative
low costs. These coagulants dissolve readily in water and the metal ions form
hexaquo complexes, which are acidic species and give up protons. The tervalent
aluminium and Fe use the specific ion adsorption for agglomeration.
Whereas the Fe (III) Chloride reacts less sensitive to pH variation, Alum
requests highest attention on the pH value. Figure 28 point out the different
physical states of the metals by changing the pH value. Figure 29 shows the
change of Al3+ from the liquid to the solid and again to the liquid physical
state by varying the pH value. To neutralise the hydrogen atoms formed in
hydrolysis of Aluminium Sulfate or Fe (III) Chloride, some carbonate hardness
is consumed (Figure 30, Figure 31).
Figure 28 - Different physical states of Alum and Iron
Figure 29 - Al3+ - change liquid - solid
Figure 30 - Alum reaction
Figure 31 - Iron (III) Chloride reaction
To take advantage of the liquid/solid properties of the alum it was foreseen
to recycle the alum. As described in Figure 30 the added Aluminium Sulfate
should coagulate with the lignin compounds in the wastewater. The resulting
sludge would be separated in a clarifier. By changing the pH value the Alum
would be soluble. With aid of Ozone or Fenton's reagent the high concentrated
lignin solution could be oxidised and the rested Alum solution would be recycled.
For the coagulation tests with Alum, Fe (III) Chloride, PAC the aerobically
pretreated PM 2 wastewater was used. As coagulant a 10 % Aluminium Sulfate,
10 % Fe (III) Chloride and 20 % Polyaluminiumchloride solution were prepared.
Figure 32 shows the connection between the solution volume and mass of the
different solutions based on the molar weight.
Figure 32 - Solution volume and mass
In each case 100 ml samples where mixed rapidly with the coagulation solution.
In the next step a magnetic stirrer should encourage the coagulation by rotating
slowly for 10 minutes followed by a sedimentation period of 30 min. Different
concentrations of the added coagulation substance should inform about the
affectivity. Especially in case of the Alum the pH value had to be monitored.
It was corrected by adding NaOH or HCl.
Figure 33 to Figure 35 describe the results obtained.
No essential effect could be observed with Aluminium Sulfate up to a concentration
of 2000 ppm. The COD concentration was only reduced from 3000 mg COD/l to
2600 mg COD/l.
Fe (III) Chloride achieved a final COD concentration of 1,900 mg/l (original
sample 3,200 mg/l) at a solution concentration of 2000 ppm.
From the PAC experiment, no measurable results could be obtained. Excluding
faults in the series of experiments it was tried to change the pH value or
the method of mixing but the results did not improve. A closer look revealed
that the size of the ligninsulfonates (0.01 to 0.05 µm) might be an
explanation as the compounds need a certain size to make sure hitting the
regency mechanism between the trivalent Fe or aluminium and the dissolved
substances.
Usually, flocculation with polymers is the subsequent step to improve the
precipitation process. But in this case a further addition of some organic
flocculants did not attain a noticeable effect.
Figure 33 - Coagulation Alum
Figure 34 - Coagulation Iron (III) Chloride
Figure 35 - Coagulation PAC
Figure 91 - photo - coagulation - Alum
Figure 92 - photo - coagulation - Fe (III) Chloride
Figure 93 - photo - coagulation - PAC
Precipitation
Precipitation is the formation of insoluble substances from dissolved matter
and the chemicals added (Ullmanns, 1988).
Lime
On the one hand Ca2+ is known in wastewater treatment as additive to improve
the precipitation of the flocks. The attendance of the cations neutralise
the negative charged anions on the surface of the flocks, leads to system
instability and as a consequence to the precipitation.
On the other hand the lime as Ca(OH)2 can build a new chemical structure with
the dissolved substances. Similar to the "soap effect" the Ca(OH)2
molecule splits off the two OH- anions and the Ca2+ cation links with the
substance.
Ca(OH)2 was dissolved in 10 ml of distilled water. The reagent was mixed
with a magnetic stirrer prior to adding 90 ml of pre-treated wastewater (aerobic
biological treatment). Subsequently, 10 minutes for reaction and 30 minutes
for settlement were given. Growing of flocks could be observed. The obvious
clarification was confirmed by COD measurements.
Lime was added in concentrations of 4, 6, 8, 10, 15, 20 g/l. An evident reduction
of the COD concentration from 3,270 mg COD/l to 1,927 mg/l could only be realised
in the range of 4 to 6 g Ca(OH)2/l.
Figure 40 - Lime precipitation - aerobic biological pre-treatment [90 ml]
Further lime precipitation tests with sample volumes of 1 to 5 l and an improved
Ca(OH)2 addition resulted in higher COD reduction (up to 61 %).
Figure 41 - Lime precipitation - aerobic biological pre-treatment [1,000 ml]
With the anaerobically pre-treated PM2 effluent the COD reduction was less
efficient. The addition of 6 Ca(OH)2 g/l led to a degradation of only 25 %.
Furthermore, increasing the lime concentration could not enhance the performance.
Figure 42 - Lime precipitation -anaerobic pre-treatment
To get an idea of the efficiency of the lime solutions applied, the COD removal
from the untreated sub-streams was also tested. The results present Figure
43 and Figure 44.
An overall COD elimination of 3,800 mg/l (52 %) in the PM2 effluents up to
10,000 mg/l (46 %) in the Plug Screw Feeder filtrate stresses the potential
of lime precipitation.
Figure 43 - Results untreated wastewater
Figure 44 - Comparison of the results - raw wastewater / biological pre-treatment
Former investigations with this type of wastewater conducted by Linde had
shown similar results. The low-molecular-weight substances in this wastewater,
such as acetic acids, carbohydrates and methanol are degraded in the previous
biological treatment. The residual oxygen demanding substances, on average
50 %, consist mainly of poor bio-degradable ligninsulfonates. Figure 36 describes
the aerobic bio-treatment of pulp and paper mill wastewater as modified incineration
(Morper, 1997).
Figure 36 - Aerobic treatment
As described in Chapter 4.1.3 Analysis, the main compound lignin and the modified
ligninsulfonate contain a number of hydroxymenthyl groups, a form of primary
alcohols that can be oxidized to the correlative carbonic acids. High-molecular-weight
salts, like soaps, are adsorbed on positively charged solids under alkaline
conditions. Whereas the aerobic pre-treatment leads to the partial oxidation
of the lignin molecules, the anaerobic treatment is a more or less chemical
reduction process that does not include such an oxidation. Therefore, lime
precipitation after anaerobic treatment brought not as much COD elimination
as aerobic pre-treatment (Morper, 1997).
Lime precipitation and potential lime recovery by calcination are described
in Figure 37.
Figure 37 - Lignin purification and decolourisation
Evaluation
By transfer of the dissolved lignin from the liquid to the solid phase by
lime precipitation, a large sludge volume resulted. By lime precipitation
(4 g/l) of aerobically treated wastewater (Figure 3-44) after 30 min of settling
the sludge volume equalled 2,500 ml of the total 5,000 ml reactor volume (corresponding
to a 500 ml in a 1,000 ml vessel). In this case, the theoretical lime consumption
of 18,000 kg/d would result in very high excess sludge production that would
be difficult to handle. For compensating reasons, further utilization or reuse
of lime should be applied. Thus, sludge treatment after lime precipitation
was investigated, too.
Basis for reuse of lime is the dewatering and volume reduction of the sludge.
Standard methods represent filter press or centrifugation.
140 ml of sludge from aerobically pre-treated wastewater (applied lime concentration:
4 g/l) was centrifuged. Subsequently, the COD concentration of the clear supernatant
was determined. The residuum, 40 ml was heated for 30 minutes at 900 °C.
The results are pointed out in Figure 38.
Figure 38 - Results of centrifugation and sludge incineration
It is obvious that after centrifugation most of the precipitated ligninsulfonates
remain in the dewatered sludge.
A filtration test should reveal the potential of applying polymers for improving
the flocculation ability. Three samples of sludge with different polymer concentrations
were filtered while measuring the water flow. As flocculent served a liquid
with a concentration of 1g/l, an anionic polymer with the trade name Prästol
2540.
The use of Prästol 2540 helped to improve the sludge dewaterability.
This could be recognised from a higher flow rate at the beginning and in the
end of the testing period. In addiion, the residuals from the polymer samples
were much better detachable from the filter (1 µm).
Figure 39 - Chemical reaction Ca(OH)2
Figure 45 - Sludge filtration test
Figure 94 - photo - precipitation - Lime - aerobic
Figure 95 - photo - precipitation - Lime - anaerobic
Figure 96 - photo - precipitation - Lime - untreated substreams
Oxidation
Methods like coagulation, flocculation, precipitation, adsorption or filtration
described in chapter 4.5) separate the refractory compounds from the effluent
to the degree of the respective efficiencies. Unless they are further treated
they are stored in the respective wastewater solids.
Another method is the oxidation by powerful chemical oxidants, which can either
mineralise organic compounds to carbon dioxide and waste or render large refractory
molecules biodegradable by cracking them into low molecular weight species.
These oxidants could enhance the reduction of the COD load. Experiences with
different oxidation technologies and different substrates showed that partial
chemical oxidation of toxic wastewater compounds may increase its biodegradability
up to high levels (Chamarro, 1999).
Fenton
In some publications the Fenton's reagent treatment is described as one of
the most effective technologies to remove organic pollutants from aqueous
solutions (Chamarro, 1999).
Fenton's reagent consists of a mixture of hydrogen peroxide and Fe salts.
There are chemical mechanisms that propose hydroxyl radicals as the oxidant
species.
Figure 46 - Principal Fenton reaction
The hydrogen peroxide reacting with ferrous ions forms a strong oxidising
agent (hydroxyl radical). The main reactions, which occur in the solution
during the Fenton process are described in Figure 47.
Figure 47 - The Fenton Process - oxidisable substance
Other possible reactions can also occur:
Figure 48 - The Fenton Process - other reaction
A process of coagulation, which occurs simultaneously to oxidation, involves
the formation of hydroxy-complexes of Fe.
Figure 49 - The Fenton Process - coagulation
The products of the reactions presented above polymerise when the pH of the
solution is kept between pH 3.5 - 7. The reactions are accurately described
in Figure 50 (Szpyrkowicz, 2000).
Figure 50 - The Fenton Process - pH between 3.5 -7
In order to investigate the suitability of the Fenton's reagent for the Sari
sub-stream treatment, a test program was developed. Based on published results
varying the reaction parameter should give a hint for the optimal composition
of the reagent.
Two important factors affecting the rate of the Fenton's reaction are the
peroxide dose and the Fe concentration. The peroxide dose is important for
the degradation efficiency, whereas the Fe concentration is important for
the reaction kinetics. Also varying the residence time or the pH value should
have additional effects. The experimental program including the variation
of standard parameters (pH, reaction time, H2O2/COD and H2O2/Fe ratio) is
shown in Figure 51, S.57. For the Fenton tests aerobically pre-treated PM2
effluent was utilised.
Figure 51 - Variations in Fenton tests
Generally, the COD could be reduced by about 25 % and a slight increase in
the BOD was observed.
Only at a H2O2/Fe ratio of 2:1 a COD reduction from 3,500 mg COD/l to 2,040
mg/l proceeded. The BOD concentration rose from 55 mg/l up to 148 mg/l (170
%).
Hence, the Fenton's reaction oxidised part of the refractory COD and also
increased the biodegradability of the PM2 effluent (Figure 52).
Figure
52 - Fenton test results
Ozone
Use of ozone for the oxidative elimination of wastewater components has been
known for a long time. Normally, ozone is used in drinking water treatment
and the sterilization of air. Its effectiveness is highly depended on the
pH and is essentially based on two following mechanisms.
The so-called direct oxidation occurs under acidic conditions. The process
is fairly slow, but the conversion can be accelerated if the energy necessary
for radical formation is provided in the form of UV light.
Figure
53 - Ozone reaction - acidic conditions
Alkaline oxidation also takes place via the intermediate formation of hydroxyl
radicals (Ullmanns, 1988).
Figure
54 - Ozone reaction - alkaline conditions
Similar to the Fenton's reaction described above the produced oxygen based
free radicals which in turn can attack many substances like lignin and thus
improve the biodegradability of the wastewater compounds. This would be attained
by a partial oxidation of the ligninsulfonates and quite low ozone doses (Arcand,
1995).
Contradictory statements exist concerning the optimum ratios of O3/COD. Because
of the specific amount of ozone varying dependent on the wastewater composition
different samples were tested.
On the one hand, ozone treatment of the aerobically pre- treated wastewater
should be a potential alternative to the previously tested lime precipitation
method. On the other hand, the solution would be an additional ozone stage
after lime precipitation.
Additionally, the combination of ozone and H2O2 was considered as an alternative
because of the specifically lower costs for H2O2 as an oxidant.
Regarding the ozone treatment of wastewater, the ozone transfer from the gas to the liquid phase must be carefully selected. Ozone is generated electrochemically from oxygen and obtained a certain concentration in a gas flow, preferably oxygen.
Traditionally similar to oxygen, the ozone is by bubbles. Mostly "frits"
(Arcand, 1995) or "injectors" (Schmidt, 2000) are in use. The alternative
"bubble-free ozone transfer" is currently investigated but is not
yet realised on technical scale application (Janknecht, 2000).
As a consequence, it was decided to use a frit for the ozone transfer. To
improve the utilisation rate a long glass cylinder should extend the rising
path the bubbles. The ozone concentration was measured before the gas mix
entered the liquid and after passing the wastewater. The Ozone dissolved in
the liquid was calculated from the concentration differences and the gas flow.
Figure
55 - Ozone experiment setup - frit
First tests with aerobically and lime treated wastewater showed that it would
be impossible to continue the experiments in this way due to intensive foaming.
Tests with an injector led to the same result. The reasons for such a reaction
are surface-active substances, which are probably abundant.
For this reason an alternative way of ozone transfer had to be applied. Generally,
pure oxygen aeration by surface aerators avoids foaming and allows a good
transfer efficiency. It takes place in a sealed reactor. This simple method
was emulated for the ozone treatment by using a little sealed glass bottle
and a magnetic stirrer.
The Ozone/air mix was directly injected in the middle of the swirl. The suck
raised by the swirl improved the gas transfer. The existing foam at the surface
of the swirl was continuously mixed with the rest of the liquid.
Figure
56 - Ozone experiment setup -surface aeration
A sidewise transfer of the gas due to the developing decompression could obtain
an additional improvement of the transfer. But this kind of transfer was not
tested.
Figure
57 - Ozone experiment -decompression- swirl
The tests were conducted with a sample volume of 250 ml, a gas flow of 20
l/s and an ozone inlet concentration between 60 - 100 g/m3.
By varying the time of treatment an ozone transfer from 0.5 - 1.7 g O3/gCOD
could be realised.
Figure 58 and Figure 59 show the progression of the ozone transfer into the
wastewater with different compositions.
Figure 58 - Diagram - Ozone entry
Figure 59 - Tables - Ozone entry
The ozone transfer gradient varied a lot. One reason for this might be the
changing pH value over time. As described above the acidic ozone reaction
needs more time and the specific transfer is lower than under alkaline conditions.
In the beginning of the test phase, the O3 consumed was quite high. In case
of the aerobically pre-treated wastewater without pH regulation it was not
possible to enter more than 0.5 g O3/gCOD. At the same time, the pH value
decreased from pH 7 to less than pH 3.
In tests with lime pre-treated wastewater it was possible to reach a level
of 1.7 g O3/gCOD. Therefore, the pH of one aerobically pre-treated sample
was increased to pH 12 by addition of NaOH. This procedure changed the transfer
behaviour completely so that a specific ozone consumption of 0.8 g O3/gCOD
could be attained.
In case of the lime sample after ozone transfer of 80 mg the foaming tendency
could not be observed anymore.
Generally, a beginning decolourisation was observed already after a short
time.
Another effect could be investigated after treating a lime probe with 0.6
g O3/gCOD - flocculation occurred. This effect is possibly due to the formation
of calcium carbonate. After the settlement of the flocks the water was rather
clear. The effect of flocculation disappeared by raising the ozone dosage.
The addition of H2O2 caused a high specific ozone transfer, but the test
results showed that the higher ozone consumption did not increase the COD
degradation.
All results of the ozone treatment tests are shown
in Figure 60.
Figure 60 - Ozone test results
The highest COD elimination could be reached by an ozone dose of 460
g (1.7 g O3/g COD) in a lime pre-treated sample. The COD concentration decreased
from 1,400 mg/l to 50 mg/l.
But also the 360 mg COD/l of the 0.6 g O3/g COD treated lime probe were respectable.
Taking the required time for ozone transfer into account, the lowest ozone
transfer of 0.6 g O3/g COD has the highest efficiency in COD removal. Only
a short (about 8 min) Ozone retention was necessary to reach the 0.6 O3/g
COD (Figure 58). This was a strong contrast to the reaction time (about 56
min) of the 1.7 g O3/g COD. This is due to the fact that the correlation between
the COD degraded and the time used for Ozone transfer is not linear.
Looking at the BOD values before and after the treatment with ozone the presumptions
in regard to a partial oxidation were verified. This effect occurred only
at the low dosed ozone samples. The BOD of the lime treated sample increased
by an ozone transfer of 0.6 g O3/g COD from 73 mg/l to 108 mg/l (32 %).
This fact confirms the experience published that two-stage ozone -biological
treatment with lower O3 loads is better than a one-stage ozone system applying
higher loadings (Helble, 1999).
Figure
100 - photo - oxidation -Ozone - experimental setup - frit - foaming
Figure
101 - photo - oxidation -Ozone - experimental setup - swirl
Figure
102 - photo - oxidation -Ozone - precipitation 1
Figure
103 - photo - oxidation -Ozone - precipitation 2
Figure
104 - photo - oxidation -Ozone - results
Figure
105 - photo - oxidation -Ozone - results - detail
4.2.3
Physical treatment
Physical treatment of wastewater has also a long tradition in wastewater treatment.
The adsorption, mostly known in connection with activated carbon, is one way
to separate the undesirable substances from water.
Adsorption
In adsorption, dissolved substances in wastewater are attracted to the adsorbents
and adhere to a solid surface. The adsorption is attributable mainly to van
der Waals forces, particularly dipole-dipole interaction, though coulombic
forces also often play an important role (Fox, 1976).
Activated Carbon and Aluminium Oxide
Potential adsorbents for adsorptive wastewater purification include activated
carbon, activated aluminium oxide (Al2O3) and diatomaceous earth.
Characteristic properties of the adsorbents, such as specific surface area,
pore volumes, and bulk densities are collected in Figure 61.
Figure 61 - Adsorbents properties
The adsorption process can be conducted either batch wise or continuously.
Principally, they can also be subdivided into mixing and filter percolation
processes (Ullmanns, 1988).
Whereas the test with activated carbon and Al2O3 were conducted as filter
process, the test with diatomaceous earth powder was conducted with the mixing
method.
A principal test setup is shown in Figure 62.
Figure 62 - Absorption Activated Carbon / Aluminium Oxide test setup
The activated carbon and the activated aluminium oxide were first mixed
with 50 - 100 ml of distilled water because of filling the pores with water,
before they could be poured into the glass cylinders. The flow was regulated
at 10 ml/min.
The activated carbon tests were realized with the pure aerobically pre-treated
PM2 wastewater, the aerobic/lime (8 g/l) and the aerobic/lime (6 g/l) pre-treatment
samples of the PM2 effluent.
The experiment's results are shown in Figure 63 to Figure 65.
Figure 63 - Absorption Activated Carbon test result - aerobically pre- treated
sample
Figure 64 - Absorption Activated Carbon test result - aerobic / lime 8 g/l
pre-treatment
Figure 65 - Absorption Activated Carbon test result - aerobic / lime 6 g/l
pre-treatment
The first 100 ml that flew through the cylinder were not usable for measurement
due to the dilution with distilled water. Based on the assumption of a plug
flow the catch liquid was separated every 50 ml.
Even if the breaking point of the activated carbon was not reached it was
obvious that the target values could not be attained.
The highest COD elimination from 1,100 mg COD/l to 650 mg/l was observed with
the aerobic/lime (8g/l) pre-treated PM2 sample. Also the aerobic/lime (6g/l)
probe could be treated 28 %.
The COD concentration of the aerobically pre-treated wastewater amounted to
3,300 mg COD/l before and 2,800 mg COD/l after the activated carbon adsorption.
Further investigations with activated aluminium oxide were conducted exclusively
with the aerobically pre- treated PM2 effluent. A COD reduction of 16 % could
be achieved. Because of the insufficient performance compared to the activated
carbon tests, aluminium oxide was excluded as a means of physical wastewater
treatment.
Figure 66 - Absorption Aluminium Oxide test result - aerobically pre-treated
sample
The testing procedure with diatomaceous earth powder and aerobically pre-treated
wastewater was similar to the lime precipitation experiments. However, the
COD reduction was quite poor: from 3,600 mg to 3,500 mg COD/l (3 %).
Figure 67 - Absorption Diatomaceous Earth powder test result - aerobically
pre-treated sample
The different results obtained can be explained by the fact that every adsorbent
can only adsorb a special kind of substance mostly depending on the nature
of the surface (acid, alkaline or neutral).
Activated carbon has a neutral surface. Consequently, it reacts preferably
with non-ionic neutral substances like AOX (Absorbable organic halogens).
The surface of activated aluminium oxide and diatomaceous earth powder is
acidic; therefore they tend to adsorb alkaline substances.
Figure 97 - photo - adsorption - experimental setup
Figure
98 - photo - adsorption - Activated Carbon
Figure 99 - photo - adsorption - Diatomaceous Earth
