Archive for May, 2012

TRICHLORAMINE IN INDOOR POOLS

TECHNICAL FOCUS:
TRICHLORAMINE IN INDOOR POOLS
Trichloramine prevention
remains better than cure
Recreation is the first UK publication to feature new
research from Germany that is relevant to anyone involved
in managing, working in or using indoor swimming pools
This German research, published here for
the first time in the UK, provides an
important step forward in our
understanding of the relationship between
chlorine levels in pools and the production
of trichloramines. Essentially, high
combined chlorine levels do not
automatically mean high trichloramine
levels.
The key precursor for the formation of
trichloramines (nitrogen trichloride) is urea
from urine, sweat and skin cells. Best
practice remains largely unchanged: that
the concentration of urea in pool water
must be minimised.
Keys to minimising urea concentration:
● Educate pool users: prevention is better
than cure. Comprehensive pre-swim
hygiene measures including using the
toilet (elimination of the urine source)
and washing thoroughly prior to pool use
(the skin source) will help greatly.
● Remove urea by water treatment through
ozone-activated carbon treatment or
photo-oxidation.
● Reduce the urea concentration by dilution
(adding 30 litres of fresh water per pool user).
● Provide good pool hall ventilation, ideally
without re-circulation or at least 30 per
cent fresh air.
BEST PRACTICE REMAINS LARGELY UNCHANGED
solubility, it readily escapes from swimming and
bathing pool water and may consequently accumulate
in the air of indoor pools and then lead to breathing
problems and eye irritations. The irritating effects
are similar to those of chlorine gas [2].
Belgian researchers hypothesised that the
exposure of schoolchildren to trichloramine during
visits to indoor chlorinated swimming pools
adversely affected the lung epithelium permeability
of the children and could lead to an increased risk
of developing asthma [3].English scientists reported
asthma symptoms in lifeguards and swimming
teachers caused by chloramines [4].More recent
studies corroborate the above hypothesis [5, 6].
In summer 1999, the German Federal
Environmental Agency, as a precaution, started to
measure trichloramine in the air of indoor pools as
part of scientific investigations into the formation
and minimisation of undesirable by-products of
swimming and bathing pool water chlorination.
The object was to obtain initial information as to
whether and to what extent the air in German
indoor swimming pools is contaminated by this
compound.No such information was available for
German indoor pools at the time.
The following is a report on the formation and
properties of trichloramine and its analysis. First
measurement results are presented and discussed.
2. Urea and formation of trichloramine in
pool water
Considerable amounts of urea are introduced to
swimming and bathing pool water by pool users.
The sources are the skin, urine and sweat.Urea is
the main final product of the protein metabolism
of humans. About 90 per cent is excreted via the
kidneys (urine), the remainder via sweat and
intestinal secretions. It also forms during skin
hornification.
Urea is a chemical compound with the following
formula:H2N-CO-NH2. In its pure form, it forms
colourless and odourless crystals which are readily
soluble in water. The presence of urea in
chlorinated pool water leads to the formation of
trichloramine.
Urea sources: skin, urine and sweat
The skin is the largest organ of the human body,
with a surface area of approximately 1.5 to 2m2.
Urea is a product of the degradation of the amino
acid arginine during skin hornification [7]. It
belongs to the natural factors that keep the skin
moist.
The urea content in the horny layer (stratum
corneum) of healthy skin is about 8 μg per cm2 of
skin surface, for both men and women. 2m2 of
skin surface would thus contain about 0.16 g of
urea. Pool water readily removes water-soluble
organic and inorganic constituents including urea
from the skin of pool users.Assuming that all urea
in the stratum corneum is fully washed into pool
water in this way, then 1,000 pool users would
release about 160g of urea into pool water.
Thorough washing and showering by pool users
prior to pool use removes about 75 to 97 per cent
of the urea contained in the stratum corneum and
is thus a very effective way to prevent urea input
into pool water (Figure 1). Substantial amounts
of urea and other nitrogen compounds may also
be introduced into pool water through urine and
sweat.Table 1 lists average concentrations of urea
in urine, sweat and the horny layer of the
epidermis.
Different figures are given in the literature as
regards urine and sweat input to pool water [9 –
13].Assuming a urine input of 35 ml per pool user
as determined by Gunkel and Jessen [9], the input
of urea to pool water would be about 0.8g per pool
user. The amount of sweat released to pool water
per pool user depends on many factors, such as
water temperature, air humidity, physical
condition and activity of the pool user. The expert
literature indicates that an active swimmer, for
example, may excrete up to one litre of sweat per
hour [14].Urea input with one litre of sweat would
amount to about 1.5g per pool user and hour.
Trichloramine formation mechanism
In the scientific literature, the mechanism of
trichloramine formation from urea is discussed
from three different directions:
• Enzymatic degradation of urea, by the enzyme
urease which is contained in various bacteria, to
ammonia or ammonium, and reaction of the latter
with free chlorine to trichloramine.According to
Jessen and Gunkel [13], this process does not occur
in chlorinated pool water;
• Hydrolysis (cleavage by the action of water) of
urea,with formation of ammonia or ammonium,
and subsequent reaction with free chlorine to
trichloramine. This occurs only at temperatures
of more than 65C and is not, therefore, relevant to
pool water;
• The decisive mechanism for the formation of
trichloramine in pool water is the step-by-step
March 2006 recreation ● 31
TECHNICAL FOCUS:
TRICHLORAMINE IN INDOOR POOLS
Figure 1: Influence of washing on the urea content in the stratum corneum, after (8).
Table 1: Average concentration of urea in urine,
sweat and the horny layer of the epidermis.
0%
5%
10%
15%
20%
25%
30%
Shower gel + water
Urea content in skin after washing
Water
40%
35% Test person A
Test person B
Test person C
Test person D
Urine
21.9g/l
Sweat
1.5g/l
Skin
8 μg/cm_
UREA CONCENTRATION
Table 3: Henry’s law constants (H).
COMPOUND H
Hypochloric acid 0.069
Monochloramine 0.45
Dichloramine 1.52
Trichloramine 435
1) Eichelsdörfer et al. [16] ; 2) Spon [17]
Table 2: Properties of chlorine and chloramines.
Free chlorine Mono Dichloramine Trichloramine
-chloramine
Mg/l
No eye irritation in rabbits1) 0-8 0-2 No data No data
Distinct eye irritation 30 4 No data No data
in rabbits1)
Odour and taste threshold2) 20 5 0.8 0.02
NCl2
O = C + 2 HOCl 2 NCl3 + CO2 + H2O
NCl2
1,1,3,3- hypochloric acid trichloramine
tetrachlorourea ➞ ➞
reaction of the urea introduced by pool users with
free chlorine to 1,1,3,3-tetrachlorourea and finally
to trichloramine, as described in the literature [15].
Properties of trichloramine
Trichloramine is an undesirable by-product of
disinfection,which has a strong irritating effect on
the eyes, nose, throat and bronchial tubes. Its
odour is similar to that of chlorine. The odour and
taste threshold in water is very low, at 0.02 mg/l.
A threshold concentration for eye irritation caused
by the presence of trichloramine in pool water has
not been established to date.
Eichelsdörfer et al. [16] demonstrated for free
chlorine and monochloramine that distinct eye
irritation in rabbits does not start at a
concentration lower than about 30 mg/l and 4
mg/l, respectively. The value for trichloramine
ought to be markedly lower. Table 2 summarises
the literature data.
Previously, it had long been assumed that
trichloramine forms only at a pH less than or equal
to 4.4. However, this view has had to be revised;
trichloramine is also formed at higher pH values
such as occur in pool water, and is rather stable
under such conditions. Investigations have shown,
for example, that a diluted aqueous trichloramine
solution has a half-life of 218 minutes at a pH of
7. This means that 50 per cent of the substance
decomposes in water during that time [18]. For
example, when the trichloramine concentration
in pool water is 0.1 mg/l, then it would be 0.05
mg/l after 218 minutes, if one ignores gaseous
emissions of the compound to air.
The outgassing behaviour of a substance
dissolved in pool water can be estimated using the
air/water partition coefficient (= Henry’s law
constant, H). The lower the Henry’s law constant,
the more soluble is the substance in pool water.
The higher its Henry’s law constant, the more
readily it escapes from pool water to air. The
Henry’s law constants of mono-, di- and
trichloramine and hypochloric acid have been
determined experimentally by Holzwarth et al.
[19] (Table 3).
The H values show that trichloramine escapes
from pool water 966 times faster than
monochloramine and 286 times faster than
dichloramine. It ‘feels’ 435 times ‘more
comfortable’ in indoor pool air than in pool water.
This and its odour and taste threshold (Table 2)
are the main reasons for the typical chlorine-like
smell in swimming pool halls. The H value of
dichloramine is only of theoretical interest, as the
compound is not stable and decomposes very
quickly in pool water [20].
Water attractions such as waterslides, water
geysers, flood showers and water fountains
accelerate the release of trichloramine to air.
Comparing, for example, the outgassing behaviour
of trichloramine to that of chloroform, which
belongs to the substance group of
trihalomethanes, trichloramine escapes from pool
water three times faster than that substance.
MEASUREMENT OF TRICHLORAMINE
3.1 Measurement in pool water
There is currently no simple on-site method for
the selective determination of trichloramine in
pool water. One laboratory method to reliably
differentiate between and quantify the various
inorganic chloramines – mono-, di- and
trichloramine – is membrane introduction mass
spectrometry (MIMS), whose use remains
reserved to specialised water analysis laboratories.
The detection limit for trichloramine is reported
to be 0.06 mg/l [21].
4.2 Measurement in the air of indoor
swimming pools
The German Federal Environmental Agency,
Department for Drinking and Swimming Pool
Water Hygiene, began measuring trichloramine
in the air of indoor swimming pools in summer
1999, for precautionary and the following other
reasons: no data whatsoever was available on
trichloramine concentrations in the air of German
swimming pool halls; the French INRS (Institut
National de Recherche et de Sécurité) has
published a validated method for the
determination of trichloramine in air [22], which
was adopted by the Federal Environmental Agency
to ensure comparability with INRS measurement
data and which is, to this day, the only existing
method for determination of trichloramine in air;
a health-based guideline value of ≤ 0.50 mg/m3
has been proposed in France for trichloramine in
indoor pool air [2, 22]. This value can be used as
a basis for assessment of the measurement results.
The principle of the analytical method is shown
in the flow chart, left.
5. Selected results
Table 4 presents selected measurement results on
trichloramine in the air of indoor swimming pools
for different pool types and compares them with
32 ● recreation March 2006
TECHNICAL FOCUS:
TRICHLORAMINE IN INDOOR POOLS
Table 5: Influence of air renewal on trichloramine concentration in indoor pool air
Flow chart: the principle of the analytical method.
Contribution of fresh air to Trichloramine Chloramines (expressed as
ingoing air mass flow in air combined chlorine) in pool water
% mg/m3 mg/l
0 0.52 0.15
30 0.37 0.15
1) according to INRS [22] ; 2) according to the German standard DIN 19643-1 [23]
Table 4: Trichloramine concentrations in the air of indoor swimming pools and corresponding
concentrations of combined chlorine in pool water.
Pool type Trichloramine Chloramines (as combined chlorine)
in indoor pool air in pool water
mg/m3 mg/l
Leisure 0.13 0.07
Leisure 0.16 0.13
Leisure 0.37 0.80
Leisure 2.2 0.12
Conventional 18.8 0.25
Hydrotherapy 0.19 0.01
Hydrotherapy 0.14 0.05
Exercise pool 0.05 0.03
Guideline values 0.501) 0.202)
CONVERSION:
Chloride concentration
per sample volume
➞NCl3 concentration
per m3 of air
➞ ➞ ➞ ➞
INDOOR
POOL AIR
(containing
NCl3, NH2Cl,
HOCl, etc.)
Selective
isolation of NCl3 by
trapping soluble
chlorine compounds
in an adsorption tube
Concentration of
NCl3 and chemical
transformation to
chloride in a special
treated filter cassette
Extraction
of chloride and
determination
by ion
chromatography
‘Water attractions such as
waterslides, water geysers,
flood showers and fountains
accelerate the release of
trichloramine to the air’
the corresponding measurement data for
combined chlorine in pool water.
The values in Table 4 show that measured
trichloramine concentrations in the air of indoor
swimming pools do not correlate with the values
for combined chlorine. There may be cases where
the concentration of combined chlorine in pool
water, at 0.80 mg/l, exceeds by far the upper value
of 0.20 mg/l recommended by the German
standard DIN 19643-1 while the trichloramine
concentration in the indoor pool air, at 0.37
mg/m3, is below the recommended guideline value
of 0.50 mg/m3.
Conversely, there are cases where concentrations
of combined chlorine in pool water comply with
(0.12 mg/l) or are just slightly above (0.25 mg/l)
the recommended upper value while the
corresponding results for the trichloramine
concentration in indoor pool air (2.2 and 18.8
mg/m3) exceed the guideline value, in the one case,
by a substantial amount. This means that a DINcompliant
concentration of combined chlorine in
pool water is not automatically linked with
trichloramine concentrations in the indoor pool
air that are safe for human health.
For this reason, care must be taken to ensure
that the ventilation system is designed so that
during pool operating hours the proportion of
fresh air fed to the air circulating in the pool hall
is adjusted to the pool capacity utilisation rate, as
prescribed by the technical rule VDI 2089-1 [24].
When the pool is used to maximum capacity (for
example,with very high bather loads and with all
water attractions switched on) the proportion of
fresh air should be at least 30 per cent of the
ingoing air mass flow.An example of the influence
of air renewal via the contribution of fresh air to
the ingoing air mass flow is presented in Table 5.
While the upper value in DIN 19643-1 for
combined chlorine in pool water is complied with,
at 0.15 mg/l, trichloramine can build up in the
indoor pool air to exceed the guideline value of
0.50 mg/m3 if there is no air renewal by a defined
proportion of fresh air (no dilution effect).
6. Discussion and conclusions
No direct correlation exists between the
trichloramine concentration in the indoor pool
air and the corresponding value for the chemical
parameter ‘combined chlorine’ in pool water.This
is because the measurement result for this sum
parameter does not, unfortunately, indicate how
much of the total content is trichloramine. A
simple and reliable on-site method for specific
measurement of trichloramine as an individual
substance in pool water does not yet exist.
A DIN-compliant concentration of combined
chlorine in pool water is no automatic guarantee
that the trichloramine concentration in the air of
the indoor pool will be tolerable from a health
perspective. In addition, this means that, during
pool operating hours, the airborne trichloramine
should be diluted by air renewal via a defined
contribution of fresh air to the ingoing air mass
flow in accordance with the generally accepted
technical standards (VDI 2089 Blatt 1) [24]. This
will prevent trichloramine accumulating in the air
of the pool hall to an extent as to exceed the
guideline value of 0.50 mg/m3.
The concentration of urea in pool water must
be minimised, since its reaction with free chlorine
in pool water results in the formation of
trichloramine, among other substances. This may
be achieved by the following:
• With the help of pool users: by using the toilet
(elimination of the urea source ‘urine’) and
washing themselves thoroughly (elimination of
the urea source ‘skin) prior to pool use;
• Removal of urea by water treatment (e.g. ozoneactivated
carbon treatment [25] [26], photooxidation
[27]); and
• Reducing the urea concentration by dilution
(adding 30 litres of fresh water per pool user).
If these hints are observed, there need be no
concern, according to present knowledge, that
trichloramine in indoor swimming pool air may
pose a health risk.
March 2006 recreation ● 33
TECHNICAL FOCUS:
TRICHLORAMINE IN INDOOR POOLS
1) DIN EN ISO 7393-2, Publication date:
2000-04, Water quality-
Determination of free chlorine and
total chlorine – Part. 2: Colorimetric
method using N,N-diethyl-1,4-
phenylendiamin, for routine control
purposes, English version, Beuth-
Verlag Berlin.
2) Gagnaire, F., Axim, S., Bonnet, P.,
Hecht, G. and Hery, M.: Comparison of
the sensory irritation response in mice
to chlorine and nitrogen trichloride. J
Appl Toxikol 14, 405-409 (1994).
3) Bernard, A, Carbonnelle, S., Michel,
O., Higuet, S., de Burbure, C., Buchet,
J.-P., Hermans, C., Dumont, X. and
Doyle, I.: Lung hyperpermeability and
asthma prevalence in schoolchildren:
unexpected associations with the
attendance at indoor chlorinated
swimming pools. Occup Environ Med
60, 385-394 (2003).
4) Thickett, K.M., McCoach, J.S., Gerber,
J.M., Sadhra, S. and Burge, P.S.:
Occupational asthma caused by
chloramines in indoor swimmingpool
air. Eur Respir J 19, 827-832
(2002).
5) Lagerkvist, B.J., Bernard, A.,
Blomberg, A., Bergstrom, E., Forsberg,
B., Holmstrom, K., Karp, K.,
Lundstrom, N.-G., Segerstedt, B.,
Svensson, M., Nordberg, G.:
Pulmonary Epithelial Integrity in
Children – Relationship to Ambient
Ozone Exposure and Swimming Pool
Attendance. Environ Health Perspect
112 (17), 1768-1771 (2004).
6) Bernard, A., Carbonnelle, S.,
Nickmilder, M., de Burbure, C.: Noninvasive
biomarkers of pulmonary
damage and inflammation:
Application to chidren exposed to
ozone and trichoramine. Toxicol Appl
Pharmacol 206 (2), 185-190 (2005).
7) Jacobi, O.: Die Inhaltsstoffe des
normalen Stratum corneum und
Callus menschlicher Haut. Arch Derm
Forsch 240, 107-118 (1971).
8) Häntschel, D., Sauermann, G.,
Steinhart, H., Hoppe, U. and Ennen, J.:
Urea analysis of extracts from
stratum corneum and the role of
urea-supplemented cosmetics. J
Cosmet Sci 49, 155-163 (1998).
9) Gunkel, K. und Jessen, H.-J.:
Untersuchungen über den
Harnstoffeintrag in das Badewasser.
Acta hydrochim hydrobiol 14, 451-461
(1986).
10) Borneff, J.: Hygiene. Georg Thieme
Verlag Stuttgart, New York, 5.
Auflage, 213, 1991.
11) Erdinger, L., Kirsch, F. und Sonntag,
H.-G.: Kalium als ein Indikator der
anthropogenen Belastung von
Schwimmbadwasser. Zbl Hyg 200,
297-308 (1997).
12) Gunkel, K. und Jessen, H.-J.: Zur
Harnstoffproblematik im
Badewasser.
Z gesamte Hyg 34, 248-250 (1988).
13) Jessen, H.-J. und Gunkel, K.: Zur
Problematik des Urineintrags in das
Badewasser. A. B. Archiv des
Badewesens Nr. 6, 273-274 (1995).
14) Roeske, W.: Schwimmbeckenwasser.
1. Auflage, Verlag Otto Haase,
Lübeck, 1980.
15) Robson, H.L.: Chloramines. In:
Encyclopedia of Chemical
Technology, Kirk, R.; Othmer, D.F. ed.,
2nd ed., Vol. 4, 908-928, John Wiley
& Sons, New York, 1993.
16) Eichelsdörfer, D., Slovak, J., Dirnagl,
K. und Schmid, K.: Zur Reizwirkung
(Konjunctivitis) von Chlor und
Chloraminen im
Schwimmbeckenwasser. Vom
Wasser 45, 17-28 (1975).
17) Spon, R.: Do You Really Have A Free
Chlorine Residual? How to Find Out
and What You Can Do About It. RR
Spon & Associates, PO box 222,
Rescoe, IL 610973, USA, 2002.
18) Cooper, W.J., Roscher, N.M., Slifker,
R.A.: Determining free available
chlorine by DPD-colorimetric, DPDSteadifac
(colorimetric), and FACTS
procedures. Journal AWWA, 362-368
(1982).
19) Holzwarth, G., Balmer, R.G. and
Sony, L.: The fate of chlorine and
chloramines in cooling towers.
Henry’s law constants for flashoff.
Water Res 18, 1421-1427 (1984).
20) Hand, V.C. and Margerum, D.W.:
Kinetics and Mechanism of the
Decomposition of Dichloramine in
Aqueous Solution. Inorg Chem 22,
1449-1456 (1983).
21) Shang, C. and Blatchley, E.R.:
Differentiation and Quantification of
Free Chlorine and Inorganic
Chloramines in Aqueous Solution by
MIMS. Environ Sci Technol 33, 2218-
2223 (1999).
22) Héry, M., Hecht, G., Gerber, J.M.,
Gendre, J.C., Hubert, G. and
Rebuffaud, J.: Exposure to
chloramines in the atmosphere of
indoor swimming pools. Ann Occup
Hyg 39, 427-439 (1995).
23) DIN 19643-1, Publication date: 1997-
04, Treatment of the water of
swimming-pools and baths – Part 1:
General requirements, English
version, Beuth Verlag Berlin
24) Technical rule (draft) VDI 2089 Blatt
1, Publication date: 2005-03 Building
services in swimming baths – Indoor
pools, English version, Beuth Verlag
Berlin
25) Eichelsdörfer, D. und v. Harpe, T.:
Einwirkung von Ozon auf Harnstoff
im Hinblick auf die
Badewasseraufbereitung. Vom
Wasser XXXVII, 73-81 (1970).
26) Jentsch, F.: Erfahrungen mit der
Ozon-Aktivkohle-Behandlung von
Schwimmbad-Meerwasser. Zbl Bakt
Hyg, I. Abt Orig B 164, 485-491
(1977).
27) Kaas, P.: Beckenwasser-
Aufbereitung mit Photooxidation.
Paper presented at the seminar
“New Concepts and Technologies”,
Dr. Jentsch Fachberatung
Schwimmbeckenwasser, Baunatal,
16 September 2003.
REFERENCES
ABOUT THE AUTHORS:
Dr Ernst Stottmeister and
K Voigt are from the
German Federal
Environmental Agency.
Email
ernst.stottmeister@uba.de

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Chlorine and Chloramine Removal with Activated Carbon

http://www.carbonresources.com/pdf/0906Potwora.pdf

Municipalities routinely began
using chlorine to treat drinking
water starting in 1908 with
Jersey City, NJ. Its use has helped to virtually eliminate diseases
like typhoid fever, cholera and dysentery in the US and other
developed countries. Globally the World Health Organization
(WHO) estimates that 3.4 million people in underdeveloped
countries die every year from water-related diseases.
Use of chlorine in water can produce an undesirable taste;
therefore, many consumers prefer to remove it. Disinfection
byproducts (DBPs) may also unintentionally form when chlorine
and other disinfectants react with natural organic matter that is
in the water. To reduce DBP formation, many municipalities are
switching to monochloramine.
Monochloramine treatment was first used in Ottawa,
Ontario, Canada in 1916 and in Denver, CO in 1917. Use of
monochloramine took a downturn during World War II due to
ammonia shortages. Currently the US EPA estimates more than
30 percent of larger US municipalities use monochloramine.
It’s a common misperception that activated carbon removes
chlorine and monochloramine from water by adsorption.
Understanding how activated carbon
removes chlorine and monochloramine
from water is critical to the design and
operation of such systems.
Chlorine formation and
reactions
Use of chlorine is the most common
method to disinfect public water
supplies. Chlorine is a powerful germicide,
killing many disease-causing
microorganisms in drinking water, reducing
them to almost non-detectable
levels. Chlorine also eliminates bacteria,
molds and algae that may grow in
water supply systems.
US EPA’s maximum residual disinfection levels (MRDLs) are
four mg/L for chlorine; however, chlorine may cause problems
that activated carbon can help resolve. The addition of chlorine
to disinfect water is accomplished by one of three forms: chlorine
gas (Cl2), sodium hypochlorite solution (NaOCl) or dry calcium
hypochlorite, Ca (OCl)2.
The addition of any of these to water will produce
hypochlorous acid (HOCl). This disassociates into hypochlorite
ions (OCl-) to some degree. (The reaction is summarized
below).
Cl2 + H2O → HOCl + H+ + Cl–
HOCl – → H+ + OCl+
The ratio of hypochlorous acid and hypochlorite ion in water
is dependent upon pH level and, to a much lesser degree, water
temperature. The ratio of hypochlorous acid and hypochlorite ion
at various water pH and temperature is shown in Table 1.
It is important to understand the ratio of hypochlorous
acid and hypochlorite ion in water. First,
it has been estimated that hypochlorous
acid is almost 100 times more effective
for disinfection than hypochlorite ion. Secondly, activated
carbons more readily remove hypochlorous acid compared to
the hypochlorite ion.
Chlorine concentrations greater than 0.3 ppm in water can be
tasted. Activated carbon is very effective in removing free chlorine
from water. The removal mechanism employed by activated
carbon for dechlorination is not the adsorption phenomena that
occur for organic compound removal.
Dechlorination involves a chemical reaction of the activated
carbon’s surface being oxidized by chlorine. There are reactions
when hypochlorous acid and hypochlorite ion react with activated
carbon (shown below).
Carbon + HOCl → C*O + H+ + Cl–
Carbon + OCl– → C*O + Cl–
C*O represents the oxidized site of activated carbon after
reacting with chlorine; the chlorine has been reduced to chloride
ion (Cl-). These reactions occur very quickly.
Factors impacting
dechlorination
When designing an activated
carbon dechlorination system, several
process factors must be considered. If
the system is being designed for organic
removal and dechlorination, design
criteria for organic removal will override
design criteria for dechlorination.
Since organic adsorption onto activated
carbon is a slower process than
dechlorination, a system that has been
properly designed for organic removal
will work well for dechlorination.
When the design is strictly for dechlorination, consideration
must be given to any dissolved organics that may be present
in the water. These organics can reduce the capacity of carbon
for dechlorination by occupying the available sites used for
dechlorination.
Particle size of activated carbon is the most important factor
impacting effective dechlorination. The smaller the activated
carbon particles, the faster the dechlorination rate. A disadvantage
of smaller particles is greater pressure drop within the media bed
and, therefore, must be given careful consideration in the overall
system design.
A 20×50 mesh size granular activated carbon (GAC) will be
more effective than a 12×30 or 8×30 mesh GAC. Carbon block
filters are made with fine mesh powder activated carbon with
particle sizes predominantly between 50 and 325 mesh.
Carbon block filters, therefore, are very effective for
dechlorination because of their very small activated carbon
particle size. If a GAC dechlorination system was designed for
20×50 mesh GAC and it was replaced with 12×40 mesh GAC, it
Chlorine and Chloramine Removal
with Activated Carbon
By Robert Potwora
Table 1. Percentages of HOCl and OCl –
% HOCl % OCl – % HOCl % OCl –
pH 32°F 32°F 68°F 68°F
4 100.0 0.0 100.0 0.0
5 100.0 0.0 97.7 2.3
6 98.2 1.8 96.8 3.2
7 83.3 16.7 75.2 24.8
8 32.2 67.8 23.2 76.8
9 4.5 95.5 2.9 97.1
10 0.5 99.5 0.3 99.7
11 0.05 99.95 0.03 99.97
J U

Using free chlorine to disinfect, however,
can cause problems. Free chlorine can
react with naturally occurring organics in
the water, like humic and fulvic acids, to
form total trihalomethanes (TTHMs) and
haloacetic acids (HAAs).
Trihalomethanes in water are generally
composed of chloroform and, to a
lesser extent, bromodichloromethane, dibromochloromethane
and bromoform. To minimize TTHM
and HAA formation, many municipalities have
switched to alternate disinfection methods, the most
common being monochloramine.
Chloramines are formed by adding ammonia to chlorinated
water. The reactions are:
HOCl + NH3 → NH2Cl + H2O (monochloramine)
HOCl + NH2Cl → NHCl2 + H2O (dichloramine)
HOCl + NHCl2 → NCl3 + H2O (trichloramine)
The chloramine formed is dependent upon water pH. At
pH less than 4.4 trichloramine is formed. Between pH 4.4 – 6.0,
dichloramine is formed. At pH above 7, monochloramine is the
most prevalent.
Since most municipalities have a pH greater than 7, monochloramine
is the only chlormaine to be concerned about. Monochloramine
may impact taste and odor, but to a lesser extent
than chlorine. It is toxic to tropical fish and may cause anemia in
patients being treated with kidney dialysis.
Removal by activated carbon, therefore, is becoming more
common. How monochloramine is removed by activated carbon
is summarized in these reactions.
GAC + NH2Cl + H2O → NH3 + H+ + Cl- + CO*
CO* + 2NH2Cl → N2 + H2O + 2H+ + 2Cl + C
CO* represents a surface oxide on the GAC
The preferred reaction is the second one because nitrogen
and chloride are the end products. With a new bed of traditional
GAC, the first reaction occurs to some degree with ammonia
being formed. Over time with traditional GAC, the second reaction
will occur.
GAC systems designed for free-chlorine removal may need to
be retrofitted for monochloramine removal. The reaction rate for
monochloramine removal is considerably slower than removing
free chlorine using traditional GAC.
At least two to four times more EBCT will be required for
monochloramine removal with traditional GAC. Regulatory
authorities and some standards may require 10 minutes EBCT
when removing monochloramine from water for kidney dialysis.
100
90
80
70
60
50
40
30
20
10
0
Percent reduction
Carbon contact elapsed time (minutes)
0 1 2 3 4 5 6 7 8 9 10
Competitive surface-enhanced
Spartan coconut surface-enhanced
Standard coconut AC
Figure 3.
Removal of monochloramine
Information on treating water for hemodialysis may be found in
ANSI/AAMI Standard RD 62:2006, “Water Treatment Equipment
for Hemodialysis Applications.”
Surface enhanced activated carbons
To compensate for poor performance of traditional GAC
for monochloramine removal, manufacturers have developed
surfaced-enhanced activated carbons. These activated carbons
have surface reaction sites enhanced during the manufacturing
process. They are superior for monochloramine removal compared
to traditional GAC. For surface-enhanced GAC, an EBCT of three
minutes will be sufficient to remove monochloramine from water.
A coconut shell-based, surface-enhanced GAC can be compared to
a bituminous coal-based, surface-enhanced
activated carbon (Table 2).
In addition to excellent monochloramine
removal with surface-enhanced coconut
shell-based GAC, its higher iodine
number means it has superior volatile
organic chemical (VOC) capacity. It also
has lower ash content and higher hardness,
resulting in less dust.
A quick bench-scale test is used to
evaluate how well different types of activated
carbons perform for monochloramine
removal. In a beaker containing 400 mL
water and four-ppm monochloramine,
0.2 grams of pulverized activated carbon
is added.
With constant stirring, reduction in
monochloramine is monitored over time. Different types of
surface-enhanced activated carbons can be compared with a
traditional activated carbon (Figure 3). The surface-enhanced,
coconut shell-based activated carbon proved superior.
Based upon field studies for surface enhanced GAC, a
minimum EBCT of three minutes is recommended. For traditional
GAC, a minimum EBCT of 10 minutes is recommended. Using the
recommended EBCT for each type of GAC, the volume of GAC
required for various flow rates may be compared (Table 3). Surfaceenhanced
GAC costs more, but based upon the lower volume
requirements, it is cost effective compared to traditional GAC.
About the author
 Robert Potwora is Technical Director for Carbon Resources, LLC. He
has 30 years experience in the activated carbon industry and is currently
Vice Chairman of the ASTM D28 Committee on Activated Carbon.
Potwora may be reached by phone at (760) 630-5724 or by email at
robert@carbonresources.com.
About the company
 Carbon Resources, based in Oceanside, CA, is a quality supplier of
activated carbon products and services that is backed by technical support
and individualized customer service. The Carbon Resources management
team has over 85 years of experience in the activated carbon industry and
offers an unmatched line of the most diverse activated carbon products
on the market. The Sabre-series®, Spartan-series® (the surface-enhanced
coconut shell-based activated carbon used in Table 2), Guardian Adsorberseries
® and newly introduced Sentry-series® activated carbon products
are widely recognized in the industry. For more information, please visit

Chemistry Induced by Hydrodynamic Cavitation

Chemistry Induced by Hydrodynamic Cavitation
Kenneth S. Suslick,* Millan M. Mdleleni, and
Jeffrey T. Ries
Department of Chemistry
UniVersity of Illinois at Urbana-Champaign
601 South Goodwin AVenue, Urbana, Illinois 61801
ReceiVed July 1, 1997
Cavitation (the formation, growth, and implosive collapse of
gas or vapor-filled bubbles in liquids) can have substantial
chemical and physical effects. While the chemical effects of
acoustic cavitation (i.e., sonochemistry and sonoluminescence)
have been extensively investigated during recent years,1-5 little
is known about the chemical consequences of hydrodynamic
cavitation created during turbulent flow of liquids. Hydrodynamic
cavitation is observed when large pressure differentials
are generated within a moving liquid and is accompanied by a
number of physical effects, erosion being most notable from a
technological viewpoint.6,7 In contrast, reports of hydrodynamically
induced chemistry or luminescence and direct comparisons
to sonochemistry or sonoluminescence have been extremely
limited.8,9
In aqueous liquids, acoustic cavitation leads to the formation
of reactive species such as OH¥, H¥, and H2O2. These shortlived
species are capable of effecting secondary oxidation and
reduction reactions. For example, iodide can be sonochemically
oxidized to triiodide by OH¥ radicals or H2O2 produced during
cavitation. From aqueous solutions containing chlorocarbons,
Cl¥ and Cl2 are also liberated in high yields and this increases
rates of iodide oxidation.10 The rate of triiodide formation is
easily monitored spectrophotometrically. For many years, this
so-called Weissler reaction has remained the standard dosimeter
for sonochemical reactions.
The recent advent of commercially available high-pressure
jet fluidizers capable of pressure drops as high as 2 kbar and
jet velocities approaching 200 m/s has led to numerous
applications in the physical processing of liquids, for emulsification,
cell disruption, etc. Chemical consequences of such
processing, however, have received little examination. One
important exception comes from W. R. Moser and co-workers,11
who have shown that such a device can be utilized to prepare
nanostructured catalytic materials. Moser speculated that the
unusual properties of his catalysts resulted from hydrodynamic
cavitation within the fluidizer.11 We describe here conclusive
experimental evidence for chemical reactions caused by hydrodynamic
cavitation within a jet fluidizer.
In a typical run,12a 60 mL of 1.0 M KI in purified water
saturated with carbon tetrachloride was introduced at a constant
flow rate into the Microfluidizer with a liquid pressure of 1.24
kbar. The reaction solution temperature increased 10 to 12 °C
within 90 s and stabilized at the temperatures reported herein.
Aliquots (4 mL) of the processed solution were periodically
extracted from the reaction system by airtight syringes, analyzed
spectrophotometrically, and returned to the reservoir after
analysis. The rate of I3
– formation was calculated from the
change in absorbance at 353 nm ( ) 26 400 M-1 cm-1)12b as
a function of reaction time. Initial studies conducted with Arsparged
water gave relatively low rates of I3
– production;
saturation of the Ar sparged H2O with CCl4 resulted in a 20-
fold increase in I3
– production, as has been typically observed
for ultrasonic cavitation.10 This is attributed to ready formation
of Cl¥ and Cl2 from CCl4 under cavitation conditions.
The effect of upstream liquid pressure on the rate of I3

production was investigated over the range 100-1500 bar. The
reaction rate increases linearly with liquid pressure (Figure 1),
but with a threshold pressure of 150 bar. Below 150 bar of
hydrostatic pressure, no chemical reactions were observed; this
probably represents the minimum jet velocity necessary to
induce cavitation. The resistance of a turbulent flow to
cavitation is given by its cavitation number ó, as defined in eq
1:6
where pd, pu, and pv are the downstream, upstream, and vapor
pressures, respectively, and the approximation holds when pu
. pd . pv, as they do under our experimental conditions. An
increase in upstream pressure should decrease ó and increase
the number of cavitation events. This in turn should increase
the rate of I3
– formation, if the chemistry is cavitation driven,
consistent with our observations.
The conditions formed during acoustic cavitation and consequently
sonochemical rates are known to be affected both by
the polytropic ratio of the dissolved gas (i.e., ç ) Cp/Cv,) and
by the thermal conductivity of the dissolved gas. The former
parameter determines the temperature achieved during bubble
compression, and the latter is responsible for heat dissipation
from the collapsing bubble to the surrounding solution. In the
present study using Ar/He mixtures, the ç of the dissolved gas
was fixed at 1.67, while the thermal conductivity was varied
from 0.017 to 0.142 W m-1 K-1. As shown in Figure 2, the
I3
– formation rate decreases exponentially as the thermal
conductivity of the dissolved gas increases. This observation
is best explained in terms of the hot-spot model for cavitation
which suggests that the maximum temperature (Tmax) realized
(1) Suslick, K. S. MRS Bull. 1995, 20, 29
(2) Mason, T. J., Ed. AdVances in Sonochemistry; JAI Press: New York,
1990-1994; Vols. 1-3.
(3) Price, G. J., Ed. Current Trends in Sonochemistry; Royal Society of
Chemistry: Cambridge, U.K., 1992.
(4) Suslick, K. S. Science 1990, 247, 1439.
(5) Suslick, K. S., Ed. Ultrasound: Its Chemical, Physical, and Biological
Effects; VCH: New York, 1988.
(6) Knapp, R. T.; Daily, J. W.; Hammitt, F. G. CaVitation; McGraw-
Hill: Inc.: New York, 1970.
(7) (a) Young, F. R. CaVitation; McGraw-Hill: Inc.: New York, 1989.
(b) Brennen, C. E. CaVitation and Bubble Dynamics; Oxford University
Press: Oxford, U.K., 1995.
(8) Anbar, M. Science 1968, 161, 1343.
(9) (a) Verbanov, V. S.; Margulis, M. A.; Demin, S. V.; Korneev, Y.
A.; Klimenko, B. N.; Nikitin, Y. B.; Pogodaev, V. I. Russ. J. Phys. Chem.
1990, 64, 1807. (b) Margulis, M. A.; Korneev, Y. A.; Demin, S. V.;
Verbanov, V. S. Russ. J. Phys. Chem. 1994, 68, 828.
(10) (a) Weissler, A.; Cooper, N. W.; Snyder, S. J. Am. Chem. Soc. 1950,
72, 1769. (b) Ibisi, M.; Brown, B. J. Acoust. Soc. Am. 1967, 41, 568. (c)
Clark, P. R.; Hill, C. R. J. Acoust. Soc. Am. 1970, 47, 649. (d) Chendke, P.
K.; Fogler, H. S. J. Phys. Chem. 1983, 87, 1362.
(11) (a) Moser, W. R.; Marshik, B. J.; Kingsley, J.; Lemberger, M.;
Willette, R.; Chan, A.; Sunstrom, J. E.; Boye, A. J. Mater. Res. 1995, 10,
2322. (b) Moser, W. R. Personal communication.
(12) (a) Reagent grade KI and CCl4 were obtained from Aldrich Chemical
Co. and used as received. High-purity water was prepared with a Barnstead
NANOpure. Spectrophotometric measurements were obtained with a Hitachi
U3300 UV-vis double-monochromator spectrophotometer. MKS mass flow
controllers 247C were used to adjust the composition of the Ar/He sparge
gas. Hydrodynamic cavitation studies were performed with an air-driven
model M-110Y Microfluidizer from Microfluidics International Corp., 30
Ossipee Rd., Newton, MA 02164. The reaction solutions were exposed only
to stainless steel or glass. The reaction solution was first sparged with highpurity
argon or Ar/He mixtures and light-proofed to prevent CCl4 photodecomposition
and then injected into the pressurizing reservoir through a
self-sealing septum. A portion of the reaction solution was pressurized by
a large pneumatically driven pump into an interaction chamber, where two
pulsed flows were redirected at each other through jewel orifices with
velocities of 190 m/s11b controlled by a back-pressure regulator. Cavitation
can occur when there is sufficient turbulence upon liquid jet impact or when
there exists a sufficient pressure drop as the streams pass through the orifices.
High-velocity pumping is also accompanied by bulk heating of the flowing
liquid. The reaction chamber, pump, and plumbing were therefore immersed
in a thermally equilibrated water bath. The processed stream was returned
to the solution reservoir for recirculation and analysis. (b) Awtrey, A. D.;
Connick, R. E. J. Am. Chem. Soc. 1951, 73, 1842.
ó )
pd – pv
pu – pd 
pd
pu
(1)
J. Am. Chem. Soc. 1997, 119, 9303-9304 9303
S0002-7863(97)02171-9 CCC: $14.00 © 1997 American Chemical Society
in a collapsing bubble decreases linearly with increasing thermal
conductivity of the entrapped gas.13 This inverse relationship
of Tmax to the thermal conductivity of dissolved gases should
lead to an exponential decrease in the I3
– formation rate
(assuming Arrhenius behavior) with increasing thermal conductivity,
as observed.
The influence of bulk solution temperature on the I3

production rate was investigated to probe the effect of vapor
pressure. The I3
– production rates decrease sharply with
increasing temperature, as shown in Figure 3. This same
behavior has been reported for many previous acoustic cavitation
studies.10b,14,15 Figure 3 also demonstrates that the observed
rates decrease exponentially with increasing solvent vapor
pressure. This same dependence is seen in sonochemical
reactions and is attributed to the increase in polyatomic vapor
inside the bubble before collapse, which decreases ç and
cushions the collapse of the cavitating bubble.15 While it is
difficult to make direct comparisons of energy efficiency (e.g.,
moles of product per kilowatt hour of electrical energy), acoustic
cavitation provides significantly higher rates for the Weissler
reaction, at least for the specific source of hydrodynamic
cavitation tested here.
In summary, we have demonstrated that the chemical effects
of hydrodynamic cavitation and acoustic cavitation respond
identically to experimental parameters, notably the bulk temperature
and the nature of the dissolved gas. In particular, the
rates decrease with increasing solution temperature, due to the
increased solvent vapor pressure inside the bubble; increasing
solvent vapor pressure attenuates the efficacy of cavitational
collapse, the maximum temperature reached during such collapse,
and, consequently, the rates of cavitational reactions. By
reducing the adiabaticity of bubble collapse, the thermal
conductivity of the dissolved gas also has a substantial effect
on the maximum temperature achieved inside a cavitating
bubble; increased thermal conductivity decreased rates of
cavitation reactions.
Acknowledgment. This work was supported by the National
Science Foundation (CHE 94-20758) and in part by the DOE. The
Microfluidizer used in this study was provided on loan from Catalytica,
Inc., and Microfluidics International Corp. We thank Dr. David L. King
and Prof. W. R. Moser for useful discussions.
JA972171I
(13) Young, F. R. J. Acoust. Soc. Am. 1976, 60, 100.
(14) (a) Sehgal, C.; Sutherland, R. G.; Verrall, R. E. J. Phys. Chem.
1980, 84, 525. (b) Didenko, Y. T.; Nastich, D. N.; Pugach, S. P.; Polovinka,
Y. A.; Kvochka, V. I. Ultrasonics 1994, 32, 71.
(15) (a) Suslick, K. S.; Gawienowski, J. J.; Schubert, P. F.; Wang, H.
H. Ultrasonics 1984, 22, 33. (b) Suslick, K. S.; Gawienowski, J. J.; Schubert,
P. F.; Wang, H. H. J. Phys. Chem. 1983, 87, 2299.
(16) (a) Washington, C.; Davis, S. S. Int. J. Pharm. 1988, 44, 169. (b)
Lidgate, D. M.; Fu, R. C.; Fleitman, J. S. Biopharm 1989, 2, 28. (c) Lidgate,
D. M.; Trattner, T.; Shulz, R. M.; Maskiewicz, R. Pharm. Res. 1992, 9,
860.
Figure 1. Dependence of triiodide formation rate on the hydrodynamic
pressure used to induce cavitation. Conditions: 60 mL of 1 M KI
solution in CCl4-saturated water was recycled under static Ar atmosphere
at a constant reaction temperature of 12 °C.
Figure 2. Dependence of triiodide formation rate on the nature of the
dissolved gas during hydrodynamic cavitation. Conditions: 60 mL of
1 M KI solution in CCl4-saturated water was recycled under static Ar
or Ar/He atmosphere at a constant reaction temperature of 12 °C and
liquid pressure of 1.24 kbar.
Figure 3. Dependence of triiodide formation rate on bulk temperature
and on solvent vapor pressure during hydrodynamic cavitation. Conditions:
60 mL of 1 M KI solution in CCl4-saturated water was recycled
under static Ar atmosphere at liquid pressure of 1.24 kbar

Combination of hydrodynamic cavitation and chlorine dioxide for disinfection of water

he combination of hydrodynamically generated cavitation with disinfectants is a new approach for disinfection of water. The aim of this study was to investigate the inactivation of Escherichia coli at high cell densities (∼107 cells mL–1) with a combination of hydrodynamic cavitation and chlorine dioxide. A new type of hydrodynamic cavitation reactor (micro wire mesh with downstream nozzle) was developed, which consumes 30% less energy than conventionally used orifice plates at similar flow rates. Compared with sole chlorine dioxide incubation the combination of chlorine dioxide with hydrodynamically generated cavitation resulted in a reduction of 50–70% of incubation time for 99% cell inactivation at an initial chlorine dioxide concentration of 0.25 mg L–1. The final viable cell number was reduced by a factor of 100–1000 compared to sole chlorine dioxide disinfection. An energy demand of ∼0.1–0.2 kWh m–1 is necessary to induce hydrodynamic cavitation at these optimum operation conditions (cavitation number of 0.13–0.17). Higher inactivation rates, the reduction of surviving cells by orders of magnitude compared to sole chlorine dioxide inactivation or the reduction of initial chlorine dioxide concentration compensate for the additional energy demand.

Cavitation Wastewater Treatment: Proof of Concept

Sergei
Godin1
Max
Fomitchev-­‐Zamilov1,2
1
Quantum
Potential
Corporation,
State
College,
PA
16803
2
Pennsylvania
State
University,
University
Park,
PA
16802
Abstract
Wastewater
is
a
frequent
byproduct
of
farming
activity
that
presents
both
opportunities
and
problems.
The
problems
arise
from
environmental
contamination
and
greenhouse
gas
(methane)
emissions
while
opportunities
reside
in
wastewater
reuse
and
recycling.
We
propose
a
feasibility
study
focusing
on
wastewater
treatment
that
will
enable
both
the
contamination
prevention
and
the
methane-­‐free
wastewater
recycling
thus
addressing
the
important
problems
of
public
health
and
climate
change.
Wastewater
contaminants
can
be
broadly
classified
as
chemical
or
biological
in
nature.
Our
proposal
falls
into
hydrodynamic
water
treatment
category,
which
has
not
yet
been
applied
to
animal
wastewater
treatment.
We
propose
to
develop
a
device,
the
activator
pump,
which
combines
the
qualities
of
hydrodynamic
and
ultrasonic
purification
systems
and
thus
is
able
to
eliminate
most
if
not
all
contaminants
regardless
of
their
nature.
The
proposed
device
is
a
modified
version
of
a
crude
oil
cracking
pump
built
by
our
company.
The
pump
comprises
a
high-­‐RPM
perforated
rotor
designed
to
create
powerful
cavitation
with
high
acoustic
power
density.
Rapidly
rotating
rotor
creates
fluid
vortices
that
result
in
rapid
and
efficient
cavitation,
which
in
turn
eliminates
solid
contaminants,
microbial
specimen
and
radicalizes
complex
organic
compounds.
We
propose
a
water
treatment
feasibility
study
using
cavitation
pump
built
by
our
company
with
the
aim
to
discover
a
set
of
operating
parameters
leading
to
most
efficient
reduction
in
contaminants
in
treated
water.
Based
on
so
gained
knowledge
we
shall
design
and
build
an
improved
water
treatment
pump
suitable
for
commercial
use.
Intended
commercial
applications
include
cost-­‐effective
detoxification
of
wastewater
lagoons
and
for
total
wastewater
recycling.
Other
potential
applications
include
water
recycling
in
residential
septic
systems
and
treatment
of
liquid
industrial
pollutants.
Development
of
Method
and
Apparatus
for
Hydrodynamic
Resonant
Wastewater
Purification
Sergei
Godin1
Max
Fomitchev-­‐Zamilov1,2
1
Quantum
Potential
Corporation,
State
College,
PA
16803
2
Pennsylvania
State
University,
University
Park,
PA
16802
Project
Narrative
Wastewater
is
a
frequent
byproduct
of
agricultural
activity,
which
presents
both
problems
and
opportunities.
The
problems
arise
from
environmental
contamination
that
the
wastewater
cause
while
opportunities
reside
in
wastewater
recycling,
which
is
particularly
important
in
generally
arid
or
high-­‐drought
areas.
We
propose
a
research
project
focusing
on
a
low-­‐cost
method
of
wastewater
treatment
that
will
enable
both
contamination
prevention
and
wastewater
recycling.
Background
Factory
Farms

a
Source
of
Water
Pollution
and
Greenhouse
Gases
Factory
farms
are
a
source
of
severe
water
pollution
[1].
Such
farms
produce
large
amounts
of
animal
waste
that
contains
significant
amounts
of
antibiotic
and
hormone
residues,
as
well
as
numerous
pathogens
such
as
phosphorus
[2].
Typically
collected
in
wastewater
lagoons

Fig.
1

animal
waste
threatens
to
leak
into
rivers,
creeks
and
underground
water
sources
linked
to
public
water
supply
and
thus
is
a
constant
source
of
legitimate
public
concern.
Fig.
1.
Animal
waste
leakage
(left)
and
a
wastewater
lagoon
(right).
Factory
farms
are
also
a
number
one
source
of
methane
and
a
major
source
nitrous
oxide
accounting
for
65%
of
worldwide
emissions
[3].
Methane
is
20-­‐times
more
powerful
green
house
gas
than
CO2
while
nitrous
oxide
is
staggeringly
300
times
more
potent
than
carbon
dioxide
[4].
Therefore
factory
farms
besides
being
an
environmental
hazard
are
a
significant
contributor
to
the
climate
change
and
global
warming.
Clearly,
something
has
to
be
done
about
animal
waste
problem.
Existing
Approaches
to
Treating
Animal
Waste
Problem
A
typical
approach
for
animal
wastewater
treatment
is
a
collection
of
waste
in
lagoons
with
subsequent
anaerobic
decomposition
[5].
Other
conventional
approaches
include
sedimentation,
aeration,
composting,
and
mechanical
treatment
[6].
These
approaches
suffer
from
high
costs
associated
with
the
need
for
careful
management,
infrastructure
and
machinery
and
generally
do
not
eliminate
the
threat
of
public
water
supply
contamination
or
greenhouse
gasses
emissions.
Novel
approaches
to
treating
animal
waste
problem
amount
to
a
conversion
of
biomass
to
energy
or
manure
to
fuel
[7,
8]
and
are
clearly
superior
to
wastewater
lagoons
since
they
eliminate
both
greenhouse
gas
emissions
and
somewhat
the
threat
of
public
water
supply
contamination.
On
the
flipside
the
construction
and
maintenance
costs
(~$1,000,000
USD)
of
the
biomass-­‐to-­‐fuel
conversion
facilities
are
substantial
and
even
higher
than
that
of
traditional
wastewater
treatment
methods
and
therefore
largely
offset
the
immediate
economic
benefits
of
these
new
technologies.
In
principle,
wastewater
treatment
is
an
energy
problem.
Given
the
unlimited
amount
energy
water
can
simply
be
boiled
in
the
presence
of
oxygen
thus
eliminating
all
microbal
contaminants
and
oxidizing
toxic
organic
compounds.
In
practice,
however,
heating
large
volumes
of
water
is
economically
unfeasible
due
to
high
energy
costs:
e.g.
it
takes
625
KWh
of
power
(or
$44
USD
at
$0.07/kwh)
to
boil
away
1
ton
of
water.
Thus,
a
better
wastewater
treatment
technology
corresponds
to
a
more
efficient
energy
transfer
from
the
power
source
(or
the
environment)
that
does
not
result
in
undesirable
bulk
water
heating,
which
amounts
to
energy
waste.
Fortunately,
such
technologies
exist
and
cavitation
is
one
of
them.
Novel
Approach

Cavitation
Cavitation
is
the
formation
of
empty
cavities
in
a
liquid
by
high
forces
and
the
immediate
implosion
of
them
[9].
Cavitation
can
be
induced
hydrodynamically
by
creating
fast
low-­‐pressure
liquid
flows
that
result
in
boiling
of
dissolved
gasses,
or
ultrasonically
by
subjecting
liquid
to
high
power
acoustic
pressures
that
essentially
tear
the
liquid
apart
by
creating
vapor-­‐filled
cavities
(bubbles)

Fig.
2.
The
unique
feature
of
cavitation
process
is
that
the
resulting
bubbles
are
short-­‐lived
and
characterized
by
high
extreme
temperatures,
pressures
and
liquid
velocities
accompanying
bubble
collapse.
In
fact
cavitation
bubbles
can
be
viewed
as
extreme
energy
concentrators
that
focus
applied
acoustic
energy
by
as
much
as
a
factor
of
1012
[10].
Temperatures
in
excess
of
30,000K
has
been
measured
in
collapsing
bubble
cores
and
pressures
in
excess
of
1,000
of
atmospheres
have
been
inferred
[11].
What
makes
cavitation
process
so
attractive
is
that
the
extreme
conditions
are
achieved
locally
(i.e.
within
bubbles)
while
the
bulk
of
the
liquid
remains
cool.
In
other
words

cavitation
is
an
efficient
process
akin
of
‘surgical
strike’
as
opposed
to
‘carpet
bombing’
of
bulk
liquid
boiling.
Fig
2.
Creation
and
implosion
of
cavitation-­‐induced
bubbles.
Cavitation
has
been
known
for
many
years
and
for
a
long
time
has
been
considered
a
parasitic
process
as
it
results
in
loss
of
pump
/
impeller
/
propeller
efficiency
and
even
massive
damage
to
surfaces
in
contact
with
collapsing
bubbles

Fig.
3

which
is
yet
another
testament
to
the
power
of
the
energy
concentration
offered
by
the
cavitation
process.
Fig.
3.
Cavitation
damage
to
a
liquid
pump
turbine.
Because
of
its
unusual
properties
cavitation
has
become
a
subject
of
intense
research
in
the
past
decade.
New
applications
of
cavitation
include
ultrasonic
cleaning
[15]
and
algae
control
[12],
chemical
processing
or
sonochemistry
[10],
petroleum
cracking
and
upgrading
[13],
and
even
industrial
wastewater
treatment
[14].
For
the
purpose
of
wastewater
treatment
cavitation
offers
the
following
two
mechanisms
of
action:
1) Shockwaves
resulting
from
powerful
bubble
implosions
disrupt
solid
particles
and
thus,
given
sufficient
treatment
time,
completely
pulverize
solid
suspensions.
This
mechanism
is
employed
commercially
for
biological
cell
disruption
on
a
lab
scale
as
an
alternative
to
sterilization
[15,
19].
2) Molecular
radicalization
results
from
extreme
temperatures
and
pressures
found
in
collapsing
bubble
cores
[11].
Water
molecules
are
known
to
disassociate
under
the
action
of
cavitation
[10]
thus
boosting
chemical
activity
of
water
due
to
pH
factor
increase
and
formation
of
hydrogen
peroxide
(H2O2).
ORP
factor
is
also
affected
and
results
in
improved
oxidation
properties
of
heavily
cavitated
water.
Thus,
the
action
of
molecular
radicalization
can
be
summarized
as
disassociation
and
oxidation
of
complex
organic
and
synthetic
compounds
contaminating
water
as
far
as
water
treatment
process
is
concerned.
From
the
theoretical
standpoint
there
is
no
question
that
cavitation
can
be
an
effective
means
for
water
purification,
consequently
a
number
of
commercially
viable
industrial
and
residential
wastewater
treatment
schemes
have
been
proposed
[17].
In
all
of
the
proposed
schemes
the
crucial
point
is
cavitation
efficiency
and
power
consumption.
Conventional
ways
for
producing
cavitation
involve
ultrasonic
generation
of
acoustic
waves
(e.g.
using
piezoelectric
transducers
or
sonotrodes)
and
hydrodynamic
cavitation
[18]
involving
whistles
(where
water
is
passed
through
one
or
more
narrow
orifices)
are
not
particularly
efficient.
E.g.
typical
efficiency
of
ultrasonic
cavitation
is
only
1-­‐10%
and
thus
is
not
commercially
attractive.
Fortunately,
there
is
another
way
to
generate
high
power
density
of
acoustic
waves
in
cavitating
liquid,
which
involves
rotor-­‐stator
apparata
[20].
Such
machines
are
built
by
our
company
and
similar
machines
are
available
from
various
manufacturers,
such
as
Arisdyne
(USA),
Kavitus
(Ukraine),
etc.
The
advantage
of
rotor-­‐stator
machines
is
that
they
rely
on
hydrodynamically
induced
cavitation
and
violent
water
jet
formation,
which
allows
transfer
of
~80-­‐90%
of
mechanical
energy
into
acoustic
energy
of
cavitation.
The
commercially
available
machines
are
primarily
used
for
petrochemical
processing,
biodiesel
production,
fuel
emulsion
preparation
and
are
yet
to
be
applied
to
wastewater
treatment
problem,
which
is
the
next
logical
step
as
far
as
the
exploration
of
the
utility
of
these
devices
is
concerned.
Technical
Description
Our
Approach

Hydrodynamic
Cavitation
We
propose
to
explore
the
efficiency
of
hydrodynamic
cavitation
using
rotor-­‐stator
machines
built
by
our
company
to
wastewater
treatment
problem
with
the
objective
to
develop
a
system
suitable
for
animal
wastewater
processing
thus
addressing
an
important
and
pressing
need
in
the
area
of
public
safety
and
climate
change.
Background
and
Preliminary
Data
Our
company
researches
and
builds
cavitation
equipment
for
petrochemical
processing
(e.g.
heavy
oil
cracking)
and
energy
research.
A
typical
machine
is
shown
on
fig.
3.
The
overall
machine
design
is
reminiscent
of
centripetal
pump
in
which
cavitation
is
maximized
as
opposed
to
being
reduced.
The
machine
shown
on
fig.
3
is
driven
by
a
50
HP
motor
and
has
a
maximum
throughput
of
20-­‐40
gallons
per
minute
(GPM).
Fig.
3.
Cavitation
rotor-­‐stator
machine
built
by
our
company.
The
internal
components
of
our
rotor-­‐stator
design
are
shown
on
fig.
4.
Fig.
4.
The
inside
view
of
rotor-­‐stator
cavitation
machine
built
by
our
companies.
The
machine
cover
is
removed
showing
the
impeller
and
the
perforated
rotor.
Water
is
forced
by
the
impeller
through
the
rotor
slots
under
high
pressure

this
is
where
cavitation
occurs.
Under
normal
conditions
the
liquid
to
be
processed
is
fed
through
the
4”
input
pipe
(center
of
the
stainless
steel
stator
enclosure,
fig.
3)
and
outputted
through
the
2”
exit
pipe
at
the
top
of
the
stator
enclosure.
The
machine
is
fully
capable
of
processing
viscous
substances
such
as
heavy
bitumen
and
does
not
choke
on
occasional
garbage.
Thus
feeding
animal
wastewater
or
any
kind
of
liquid
substance
with
substantial
solid
content
does
not
present
a
problem.
While
we
have
initially
designed
the
machine
to
work
on
hydrocarbons
we
have
performed
the
following
measurements
on
water,
these
constitute
our
preliminary
‘seed
data’
that
in
our
view
justifies
an
expanded
investigation
into
wastewater
treatment
problem:
1) Typical
stator
pressure
is
80
psi
(6
atmospheres),
pressures
in
excess
of
200
psi
are
possible
with
extended
impeller
blades
(not
shown
on
fig.
4).
High
stator
pressure
is
instrumental
in
achieving
very
strong
cavitation.
In
fact
under
120
psi
the
cavitation
is
already
10-­‐times
more
extreme
than
a
typical
lab
experiment
involving
ultrasonic
cavitation.
2) Typical
throughput

20
GPM
at
25kW
and
40
GPM
at
40kW.
3) Typical
pH
changes
from
7.5
(neutral
non-­‐processed
water)
to
6.0-­‐6.5
(acidic).
4) Solid
particle
destruction

ground
coffee
beans
were
run
through
the
machine
with
1mm
grains,
the
resulting
ground
mix
suspension
contained
such
a
small
particles
(micron
and
submicron)
that
we
could
not
identify
their
size.
5) Chemical
reaction
activation

we
have
ran
a
1-­‐mole
solution
of
potassium
iodine
(KI)
through
the
machine
and
obtained
a
solid
powder-­‐like
sediment
indicative
of
intensive
sonochemical
reactions
that
we
were
unable
to
produce
during
ultrasonic
cavitation.
6) Efficiency

the
machine
is
81%
efficient
in
converting
electric
energy
into
acoustic
energy
(the
motor
is
96%
efficient
and
the
frequency
control
unit
is
95%
efficient,
the
remaining
10%
losses
amount
to
bearing
heating,
water
pumping
power
and
hydrodynamic
losses).
7) Estimated
acoustic
power
density

1
MW/m2.
8) Microbal
specimen
elimination

for
a
test
we
have
circulated
25G
of
pond
water
through
the
machine
for
10
minutes;
at
the
end
of
the
test
we
were
not
able
to
detect
any
microorganisms
in
the
processed
water.
Feasibility
Evaluation
In
our
opinion
the
preliminary
data
warrants
further
investigation
into
the
wastewater
treatment
using
our
rotor-­‐stator
equipment
because
the
process
may
be
feasible
on
commercial
scale.
E.g.
20
GPM
processing
rate
at
25
kW
of
electric
power
results
in
20
GPM
x
3.78
L/Gal
x
60
min
/
25
kW
=
181L
of
treated
water
per
kWh,
which
is
equivalent
to
$0.38
USD
per
ton
of
wastewater
(assuming
$0.07/kWh).
This
estimate
makes
hydrodynamic
cavitation
process
over
100
more
efficient
compared
to
brute
force
water
boiling.
Based
on
the
preliminary
data,
as
a
result
of
the
cavitation
processing
we
expect
the
following
transformations
in
the
processed
water:
1) Large
solid
particles
to
be
pulverized
to
micron-­‐size;
2) Complete
elimination
of
microbal
specimen
(including
pathological
bacteria);
3) At
least
some
degree
of
complex
organic
molecule
destruction
and
oxidation
due
to
sonochemical
reactions
and
oxidation
due
to
the
increase
in
pH
and
other
factors
accompanying
intense
cavitation
bubble
collapse.
In
other
words,
cavitation
wastewater
treatment
may
be
suitable
for
eliminating
Class
A
biosolids,
including
pathogenic
bacteria
and
pharmaceuticals
(e.g.
via
increased
pH).
Possible
Commercial
Utilization
Scenarios
Assuming
that
the
detailed
feasibility
study
produces
positive
results,
we
envision
several
applications
of
the
cavitation
wastewater
treatment:
1) In
the
factory
farm
setting
the
cavitation
pump
can
be
used
to
pre-­‐treat
water/manure
mixture
discharged
into
the
lagoon
thus
making
the
lagoon
less
hazardous;
2) Lagoons
with
large
water
content
can
be
pumped
out
with
the
cavitation
pump
and
because
the
cavitation
process
eliminates
pathogenic
bacteria
and
class
A
biosolids
the
pumped-­‐out
water
can
be
reused
for
irrigation;
3) The
cavitation
water
treatment
system
can
be
combined
with
methane
production
via
anaerobic
bacteria
forming
a
closed-­‐loop
system.
While
the
latter
scenario
is
the
most
capital
intensive
in
terms
of
upfront
hardware
requirements,
it
is
the
most
lucrative
in
the
long
run.
In
this
scenario
the
wastewater
or
sludge
(regardless
of
the
solids
content)
can
be
processed
as
follows:
1) The
liquid
is
pretreated
with
the
cavitation
pump
to
eliminate
all
bacterial
specimen,
toxins
and
pulverize
solid
particles;
2) The
pretreated
water/sludge
is
deposited
in
a
tank
and
mixed
with
anaerobic
organisms
for
methane
production;
3) Due
to
elimination
of
competing
microbial
species
and
food
particle
pulverization
one
can
expect
‘explosive’
methane
production
from
the
tank;
4) The
methane
can
be
collected
in
a
tank
and
used
for
the
farm
needs
such
as
heating
and
powering
generators
with
portion
of
the
power
feeding
the
pump
and
auxiliary
equipment
necessary
to
support
the
process;
5) The
digested
solid
residue
can
be
dried
and
pelletized
and
used
for
fertilization
thus
producing
a
closed-­‐loop
and
waste-­‐free
system.
In
fact
a
similar
scenario
is
already
realized
in
pilot
projects
in
Russian
Federation
(e.g.
“Special
Technologies”,
Moscow).
But
the
equipment
is
not
yet
available
for
export
outside
the
country.
Specific
Aims
During
Phase
I
of
the
project
we
plan
to
achieve
the
following:
1) Test
our
cavitation
pump
using
actual
wastewater
samples
obtained
from
nearby
farms
and
University
Area
Joint
Authority
(sewage)
to
determine
the
degree
of
pathogen
removal
possible
with
our
existing
equipment
(the
water
samples
will
be
analyzed
at
Penn
State’s
Agricultural
Analytical
Services
Lab,
which
specializes
on
water
quality
control).
We
will
pay
special
attention
to
bacterial
and
chemical
content
as
well
as
the
solid
particle
size
before
and
after
treatment.
2) Experiment
with
various
pump
operating
modes
to
optimize
throughput
while
maintaining
satisfactory
treated
water
quality.
3) Based
on
the
data
so
obtained
design
and
build
an
improved
cavitation
pump
geared
towards
maximum
wastewater
treatment
efficiency.
The
new
pump
will
serve
as
a
commercial
device
prototype
for
Phase
II
of
the
project.
Potential
Post-­‐Applications
Given
successful
completion
of
Phases
I
and
II
of
the
project
we
intend
to
build
and
market
the
following
range
of
wastewater
treatment
equipment:
1) Cavitation
pump
for
sewer
water
treatment
for
municipal
applications
(including
human
waste);
2) Cavitation
pump
for
animal
farm
wastewater
pre-­‐treatment
and
lagoon
water
purification;
3) Closed-­‐loop
system
that
combines
cavitation
water
treatment
with
anaerobic
methane
production.
The
target
efficiency
is
better
than
$0.40/ton
of
treated
waste.
Therefore
the
application
of
the
activator
pump
to
wastewater
treatment
directly
addresses
the
clean
water
requirements
and
greenhouse
gas
emission
standards
mandated
by
the
EPA
for
the
farming
industry.
The
cavitation
pump
is
a
simple
and
low-­‐
cost
solution
to
the
problem
that
is
more
environmentally
friendly
and
easier
to
implement
then
existing
wastewater
treatment
techniques.
Potential
additional
applications
include
adaptation
of
the
hydrodynamically
induced
cavitation
technology
to
various
specific
chemical
substances
elimination
(focusing
on
most
common
liquid
pollutants),
further
development
of
crude
and
liquefied
bitumen
cracking
and
nanoparticle
production.
Satisfaction
of
the
Public
Interest
If
successful
the
activator
pump
will
address
the
following
important
Strategic
Goals:
-­‐ Sustainability
in
Rural
Farm
Economic,
achieved
by
eliminating
or
reducing
wastewater
lagoons
and
enabling
wastewater
recycling
and
reuse
-­‐ Improvement
of
Nation’s
Health,
achieved
by
eliminating
or
reducing
wastewater
lagoons
and
thus
mitigating
public
water
supply
contamination
threats
due
to
toxic
waste
leakage,
elimination
of
putrid
farm-­‐associated
orders
-­‐ Protection
of
the
Environment,
achieved
by
minimizing
hazardous
qualities
of
activator-­‐treated
wastewater,
reduction
or
elimination
of
the
wastewater
lagoons.
Experiment
Plan
To
test
water
treatment
efficiently
of
the
cavitation
pump
we
will
build
an
experimental
setup
similar
to
the
one
shown
on
fig.
5
except
that
we
be
mounted
on
a
truck.
Currently
the
cavitation
pump
is
powered
by
an
electric
motor
requiring
a
3-­‐phase
power,
which
makes
it
necessary
to
employ
a
diesel
generator
to
power
the
pump
in
the
field
conditions.
The
bypass
pipe
(see
on
fig.
5)
will
be
used
to
control
the
flow
rate
and
cavitation
time.
Fig.
5.
Cavitation
pump
water
experiment
with
a
single
barrel.
The
pump
will
move
the
wastewater
from
the
reservoir
and
back
and
we
shall
obtain
2-­‐3
samples
of
the
running
water
to
the
following
tests:
• Solids
content;
• Solids
size;
• Total
coliform
and
E.
coli
bacteria;
• pH;
• Nitrate-­‐nitrogen
content;
• Phosphate
and
sulfate
content;
• Ammonia
content;
• Total
organic
carbon
(TOC)
content;
• Hormone
content
(estrogen).
The
tests
will
be
subcontracted
to
Penn
State’s
Agricultural
Analytical
Services
Lab.
The
variables
that
we
will
control
are:
• Stator
pressure
(monitored
via
manometer);
• Flow
rate
(monitored
via
flow
meter);
• Cavitation
intensity
(monitored
acoustically
via
high
frequency
pressure
transducer
connected
to
spectrum
analyzer);
• Stator
temperature
and
pH
value
(monitored
via
pH
probe
with
RTD
thermocouple).
The
objective
of
the
trial
runs
on
various
water
samples
is
to
determine
optimal
set
of
control
variables
that
results
in
complete
elimination
of
bacteria
and
reduction
in
TOC,
nitrate,
phosphate
and
ammonia
content.
Work
Schedule
As
a
part
of
Phase
I
of
the
project
we
plan
to
do
the
following:
1) Procure
a
flatbed
truck
and
a
mobile
diesel
generator
(Week
1)
a. The
work
will
be
performed
by
Technician.
2) Mount
cavitation
pump
on
a
truck
(Week
3)
a. The
work
will
be
performed
by
PI
and
Co-­‐PI
with
the
help
of
Technician.
b. Machining
work
will
be
subcontracted
to
Butler
Machine
Shop.
3) Perform
a
series
of
field
tests
with
this
mobile
system
and
accumulate
trials
data
(Week
5)
a. Field
work
will
be
performed
by
Technician.
b. Advice
on
location
selection,
test
configuration
and
execution
will
be
given
by
Consultant.
c. Analytical
work
will
be
subcontracted
to
Penn
State’s
Agricultural
Analytical
Services
Lab.
4) Analyze
the
data
to
determine
the
most
efficient
operating
mode
and
infer
cavitation
parameters
from
it
(Week
9)
a. The
work
will
be
performed
by
PI
with
the
help
of
Consultant.
5) Using
the
so
obtained
cavitation
parameters
design
and
build
an
improved
and
scaled
down
version
of
the
pump
that
does
not
require
3-­‐phase
power
and
can
work
from
120V
(Week
10)
a. The
work
will
be
performed
by
PI
and
Co-­‐PI
with
the
help
of
Technician.
b. Machining
work
will
be
subcontracted
to
Butler
Machine
Shop.
6) Repeat
a
series
of
field
tests
to
confirm
the
efficiency
of
the
redesigned
pump
(Week
30)
a. Field
work
will
be
performed
by
Technician.
b. Advice
on
location
selection,
test
configuration
and
execution
will
be
given
by
Consultant.
c. Analytical
work
will
be
subcontracted
to
Penn
State’s
Agricultural
Analytical
Services
Lab.
7) Write
final
report
(Week
32)
a. Work
will
be
performed
by
PI.
References
[1]
M.D.
Allen,
Understanding
State
Adoptions
of
Factory
Farm
Regulation,
proc.
of
Western
Political
Science
Association,
Albuquerque,
New
Mexico,
2006.
[2]
S.
Wing,
S.
Freedman,
L.
Band,
The
Potential
Impact
of
Flooding
on
Confined
Animal
Feeding
Operations
in
Eastern
North
Carolina,
Environmental
Health
Perspectives,
110,
387-­‐390,
2002
[3]
Sources
and
Emissions:
Methane,
U.S.
Environmental
Protection
Agency,
2006,
http://epa.gov/methane/sources.html
[4]
Climate
Change:
Methane,
U.S.
Environmental
Protection
Agency,
2006,
http://www.epa.gov/methane/
[5]
R.
Zhang,
Biology
and
Engineering
of
Animal
Wastewater
Lagoons,
http://groups.ucanr.org/LNM/files/678.pdf
[6]
S.
Mukhtar,
Animal
Manure
and
Process-­‐Generated
Wastewater
Treatment,
2003,
http://www.cals.ncsu.edu/waste_mgt/natlcenter/modules/Module_5(new).doc
[7]
Biomass
Energy:
Manure
for
Fuel,
Texas
State
Energy
Conservation
Office,
2008,
http://www.seco.cpa.state.tx.us/re_biomass-­‐manure.htm
[8]
B.
Min
et
al.,
Electricity
generation
from
swine
wastewater
using
microbial
fuel
cells,
Water
Research,
39,
4961–4968,
2005
[9]
C.
Brennen,
Cavitation
and
Bubble
Dynamics,
Oxford
University
Press,
1995
[10]
M.
Margulis,
Sonochemistry
and
Cavitation,
Gordon
and
Breach,
1993
[11]
D.
Flanigan,
K.
Suslick,
Internally
confined
plasma
in
an
imploding
bubble,
Nature
Physics,
6,
pp.
598-­‐601,
2010
[12]
LG
Sound,
http://www.lgsonic.com
[13]
New
Technologies
2000,
LLC,
http://www.NewTech2000.ru/index_eng.php,
2008
[14]
A.
Chakinala,
Industrial
wastewater
treatment
using
hydrodynamic
cavitation
and
heterogeneous
advanced
Fenton
processing,
Chemical
Engineering,
152,
498-­‐502,
2009
[15]
L.
Azar,
Cavitation
in
Ultrasonic
Cleaning
and
Cell
Disruption,
Controlled
Environments,
2009,
http://www.megasonics.com/Cavitation.pdf
[16]
Y.
Benito
et
al.,
Hydrodynamic
Cavitation
as
low-­‐cost
OP
for
wastewater
treatment:
preliminary
results
and
a
new
design
approach,
WIT
Transactions
on
Ecology
and
the
Environment,
80,
p.
495
[17]
P.
Gogate,
A.
Pandit,
A
review
of
imperative
technologies
for
wastewater
treatment
I:
oxidation
technologies
at
ambient
conditions,
Advances
in
Environmental
Research,
8,
2004,
pp.
501-­‐551
[18]
P.
Gogate,
Cavitation:
an
auxiliary
technique
in
wastewater
treatment
schemes,
Advances
in
Environmental
Research,
6,
2002,
pp.
335-­‐358
[19]
P.
Gogate,
Hydrodynamic
Cavitation
for
Food
and
Water
Processing,
Food
and
Bioprocess
Technology,
4,
6,
pp.
996-­‐1011
[20]
M.
Promtov,
Pulsation
Apparata
of
Rotor
Type:
Theory
and
Practice,
Mashinostroyeniye,
2001
(In
Russian)
Facilities
The
Hanger
Our
company
has
a
800
sq.
ft.
lab
at
Penn
Eagle
Industrial
Park
at
100
Rolling
Ridge
Drive,
Bellefonte,
PA
suitable
to
conduct
the
proposed
investigation.
Water
Testing
We
plan
on
contracting
Penn
State
Agricultural
Analytical
Services
Lab
for
water
testing:
http://www.aasl.psu.edu/waterprogram_main.html.
The
lab
provides
services
on
commercial
basis
for
local
businesses
and
farmers.
Machining
We
plan
to
subcontract
machining
work
to
Butler
Machine
Shop,
Bellefonte,
PA.
Equipment
We
plan
to
acquire
the
following
major
equipment
necessary
for
conducting
of
the
research
under
this
proposal:
1) Used
flatbed
truck,
Ford
F-­‐150
or
similar:
$20,000
2) Towable
diesel
generator,
Magnum
40
kW
or
similar:
$15,000
The
company
has
all
other
equipment
necessary
to
conduct
the
work.
Budget
Justification
Diesel
Generator
Is
necessary
to
provide
40
kW
of
3-­‐phase
power
for
the
cavitation
pump
during
the
field
trials.
Flatbed
Truck
Is
necessary
to
transport
the
cavitation
pump
and
the
generator
to
field
locations.
Lab
Testing
$150-­‐200/test
(Penn
State
Agricultural
Analytical
Services
Lab)
Materials
for
Machining
AC
motor,
stainless
steel:
$8,000
Max
Fomitchev-­‐Zamilov,
Ph.D.
(coPI)
President,
Quantum
Potential
Corporation
Assistant
Professor,
Pennsylvania
State
University
Biographical
Sketch
Dr.
Fomitchev-­‐Zamilov
is
the
director
of
and
the
vision
behind
the
Quantum
Potential
Corporation.
The
mission
of
the
company
is
identification,
analysis
and
exploration
of
promising
yet
neglected
lines
of
research
with
the
focus
on
high-­‐
risk/high-­‐payoff
projects
(inline
with
latest
federal
initiatives).
During
the
past
decade
Quantum
Potential
has
amassed
a
vast
portfolio
of
research,
sponsored
and
launched
a
number
of
research
project
and
obtained
patent-­‐pending
commercializable
results.
Currently
Quantum
Potential
is
actively
pursuing
cooperation
with
NASA,
NIH
and
DoE.
Main
projects
include
cavitation
hydrocarbon
processing
with
applications
to
heavy
oil
upgrading
and
electromagnetic
equipment
for
medical
treatment.
Dr.
Fomitchev-­‐Zamilov’s
role
is
that
of
a
physicist,
engineer,
and
administrator.
Having
cultivated
broad
encyclopedic
knowledge
from
various
disciplines
in
science
Dr.
Fomitchev-­‐Zamilov
is
working
on
pursuing
collaboration
between
like-­‐minded
individuals
and
organizations
in
order
to
facilitate
the
nucleation
of
the
next
technological
breakthrough.
Of
particular
relevance
to
this
proposal
is
Dr.
Fomitchev-­‐Zamilov’s
experience
with
cavitation
equipment
and
optimal-­‐lag
ultrasonic
pulse
shaping.
Education
2000-­‐2001,
Moscow
Institute
of
Electronic
Engineering,
Ph.D.,
Computer
Engineering
1997-­‐1998,
The
University
of
Tulsa,
Ph.D.
Candidate,
Computer
Science
1992-­‐1997,
Moscow
Institute
of
Electronic
Engineering,
M.S.,
Computer
Technology
Positions
2006-­‐present,
Pennsylvania
State
University,
Assistant
Professor
of
Computer
Science
2002-­‐present,
Quantum
Potential
Corporation,
Director
Patents
&
Publications
Dr.
Fomitchev-­‐Zamilov
has
authored
two
books
and
dozens
of
papers
and
articles
in
the
field
of
computer
science,
engineering
and
physics;
he
also
holds
two
patents.
Relevant
Publications
Fomitchev,
M.I.,
US6167758,
Ultrasound
Imaging
Device
that
Uses
Optimal
Lag
Pulse
Shaping
Filters,
issued
01/02/2001.
Fomitchev
et
al.,
Ultrasonic
Pulse
Shaping
with
Optimal
Lag
Filters,
International
Journal
of
Imaging
Systems
and
Technology,
10,
5,
pp.
397-­‐403,
1999
Sergei
Godin
(PI)
R&D
Director,
Quantum
Potential
Corporation
Biographical
Sketch
Mr.
Godin
is
an
experienced
practitioner
and
an
exceptional
experimentalist.
He
is
an
expert
in
electrical
engineering,
digital
/
analog
electronics,
measurement
devices
and
experimental
setup
design.
Prior
to
joining
Quantum
Potential
Mr.
Godin
has
worked
as
an
engineer
at
the
Central
Research
Institute
for
Communications
(Moscow),
then
as
a
research
associate
at
IMASH
(Moscow)
and
for
the
following
12
years
as
a
research
associate
at
the
Institute
for
High
Temperatures
(IHT)
of
the
Russian
Academy
of
Sciences.
During
his
tenure
at
IHT
Mr.
Godin
was
a
key
investigator
in
a
number
of
research
projects
focused
on
sonoluminescence,
cavitation,
plasma
discharges,
and
nuclear
fusion.
Because
of
his
prior
experience
with
hydrodynamic
cavitation
and
oil
cracking
pumps
Mr.
Godin
is
the
best
qualified
person
to
lead
the
project.
Mr.
Godin
has
a
valuable
experience
of
research
commercialization
and
has
a
knack
for
discovering
multiple
practical
applications
of
scientific
ideas.
He
leads
a
diverse
group
of
cross-­‐disciplinary
researchers.
Besides
his
duties
at
Quantum
Potential
Mr.
Godin
servers
as
a
consultant
on
a
oil
cracking
research
project
for
a
large
Russian
oil
and
gas
company.
Mr.
Godin
has
co-­‐authored
a
book
on
fundamental
physics,
numerous
research
papers
and
holds
several
patents.
Education
1988-­‐1989,
Moscow
State
University,
MechMat,
Ph.D.
Candidate
1982-­‐1983,
Moscow
Institute
of
Radio-­‐engineering
and
Automation,
Certificate
of
Accomplishment
in
Signal
Processing
1976-­‐1981,
Moscow
Institute
of
Communications
and
Informatics,
M.S.,
Electrical
Engineering
Positions
1996-­‐2008,
Institute
for
High
Temperatures
of
Russian
Acad.
of
Sci.,
Research
Associate
2010-­‐present,
Quantum
Potential
Corporation,
Research
Associate
Relevant
Publications
1. Karimov,
A.R.,
Godin,
S.M.,
Coupled
radial–azimuthal
oscillations
in
twirling
cylindrical
plasmas,
Physica
Scripta,
80,
3,
2009
2. Godin,
S.M.,
Botvinsly,
V.
V.,
Measurements
of
displacement
current
with
fammeter,
Radiotechnology
&
Electronics,
54,
9,
2009,
1049-­‐1152
3. Godin,
S.M.,
Rodionov,
B.U.,
Savvatimova,
I.B.,
Inspection
method
to
check
quality
of
nuclear
transmutation
media,
The
13th
International
Conference
on
Condensed
Matter
Nuclear
Science,
2007,
Dagomys,
Russia
4. Roschin,
V.V.,
Godin,
S.M.,
Orbiting
Multi-­‐Rotor
Homopolar
System,
US
Patent
#6,822,361,
2004
5. Klimov
et
al.,
On
the
possibility
of
electrostatic
relativistic
dimano,
Radiotechnology
and
Electronics,
49,
11,
2004,
1237-­‐1243
6. Klimov
et
al.,
The
use
of
the
relativistic
effect
for
obtaining
negative
permittivity,
International
Conference
on
Antenna
Theory
and
Techniques,
Sevastopol,
Ukraine,
vol.
1,
2003,
171

172
7. Klimov
et
al.,
The
model
of
creation
of
rotating
stationary
electromagnetic
formations
in
vacuum,
International
Conference
on
Antenna
Theory
and
Techniques,
Sevastopol,
Ukraine,
vol.
1,
2003,
173

177
8. Zolotarev,
V.F.,
Roschin,
V.V.,
Godin,
S.M.,
On
the
Structure
of
Space-­‐Time
and
Certain
Fundamental
Interactions,
Moscow,
2000,
ISBN
5862030875
Bruce
Logan,
Ph.D.
(Consultant)
Kappe
Professor
of
Environmental
Engineering,
Pennsylvania
State
University
Biographical
Sketch
Dr.
Logan
is
an
expert
in
water
treatment
and
environmental
engineering.
He
is
engaged
in
the
development
of
new
bioelectrochemical
technologies
for
achieving
an
energy
sustainable
water
infrastructure.
Logan
and
his
collaborators
have:
invented
a
method
for
sustainable
hydrogen
production
using
microbial
electrolysis
cells
(MECs);
invented
a
method
for
water
desalination
that
does
not
require
electrical
energy
from
the
grid
or
high
pressures
called
microbial
desalination
cells;
improved
direct
bioelectricity
generation
by
several
orders
of
magnitude
in
microbial
fuel
cells
(MFCs).
Other
research
has
included
the
discovery
of
how
large
aggregates
form
in
the
ocean,
called
marine
snow,
that
can
help
to
sequester
carbon
to
deep
sediments;
and
molecular
and
nanoscale
techniques
to
study
particle
dynamics
and
microbial
adhesion
in
engineered
and
natural
systems;
microbial
adhesion
and
transport.
Education
1986
Ph.D.
in
Environmental
Engineering,
University
of
California,
Berkeley,
CA
1980
M.S.
in
Environmental
Engineering,
Rensselaer
Polytechnic
Institute,
Troy,
NY
1979
B.S.
in
Chemical
Engineering,
Rensselaer
Polytechnic
Institute,
Troy,
NY
Appointments
Home Dept: Kappe Professor of Environmental Engineering, Dept. of Civil and
Environmental Engineering
Chemical Engineering; Nuclear & Mechanical Engineering
Director: Engineering Energy & Environmental Institute
Director: Hydrogen Energy (H2E) Center
Dr. Logan has numerous patents, publications, and awards. For complete list of his
accomplishments please see his CV at Complete
CV:
http://www.engr.psu.edu/ce/enve/logan/Logan_CV.pdf

VOLATILISATION PROCESSES IN WASTEWATER TREATMENT PLANTS AS A SOURCE OF POTENTIAL EXPOSURE TO VOCs

VOLATILISATION PROCESSES IN WASTEWATER
TREATMENT PLANTS AS A SOURCE OF POTENTIAL
EXPOSURE TO VOCs
Alexander P. Bianchi*f and Mark S. Varneyf
•Industrial Hygiene Department, Exxon Chemical Co., Cadland Rd, Hythe, Southampton SO45 6NP,
U.K., and tDepartment of Oceanographic Sciences, Oceanography Centre, University of Southampton,
Empress Dock, Southampton SO40 4WH, U.K.
(Received 17 October 1996)
Abstract—The results of a survey estimating volatilisation rates (as gas exchange constants and
flux rates) of a range of hazardous alkane, aromatic, organohalogen, organosulphide, ketone and
alcohol volatile organic compounds (VOCs) from a water bay are presented. More than 73
experimental test runs were earned out over a year under varying seasonal conditions to measure
the simultaneous concentrations of VOCs in air and water phases. The data were processed
employing new advances in ‘surface renewal’ volatilisation models. Based on the experimental
data, estimates of theoretical equilibrium constants, gas exchange constants and flux rates for each
VOC were made. Within the ranges of concentration which may reasonably be encountered by
process workers, the fluxes for a broad range of VOCs ranged from approximately 0.04x 10~8 to
9.0×10″‘ g cm”2 h~’. The results were used to make predictions about the relationship between
volatilisation processes and their implications for occupational hygiene nsk assessment. © 1997
British Occupational Hygiene Society. Published by Elsevier Science Ltd
INTRODUCTION
Emissions of volatile organic compounds (VOCs) from wastewaters in municipal
sewage treatment plants, surface impoundments, waste lagoons, industrial wastewaters
and drainage systems are often overlooked as sources of exposure to
hazardous substances. In many cases, the toxic effects of a broad range of VOCs
arising from such sources may have significant adverse consequences for public
health (Pellizzari, 1982; Namkung and Rittman, 1987; Ciccioli, 1993; Wallace, 1993)
and for wastewater plant workers within an industrial setting (de Mik, 1993). The
range of volatile chemical substances passing through wastewater treatment plants
are often unknown, which creates problems when attempting to measure the VOC
contribution to airborne environments or assess exposure implications for operating
personnel (Deacon, 1977; Dix, 1981; Berrafato and Wadden, 1986; Maugh, 1987;
GESAMP, 1991; Haz. Sub., 1993; de Mik, 1993). Up to 40% of the organic loading
within an industrial or municipal wastewater plant may be volatile (for example
VOCs with boiling point: <350°C; vapour pressure, 0.1-500 mm Hg at 20°C)
depending on the nature of materials flushed to sewer (Knap et al., 1979; Clark,
1986). Ecotoxicological impacts within the immediate receiving environment are
also closely associated with the effects of toxic vapour and liquid-phase VOCs in
effluent (Verscheuren, 1983; Stagg, 1986).
^Author to whom correspondence should be addressed.
437
Downloaded from http://annhyg.oxfordjournals.org/ by guest on May 11, 2012
438 A. P. Bianchi and M. S. Varney
Volatilisation has long been recognised as a mechanism whereby organic
compounds with appropriate physicochemical characteristics (that is aqueous
solubility and concentration, vapour pressure and Henry's Law constant) transfer
across the water/air interface from the 'aqueous' compartment to the 'atmosphere'
compartment (Kotzias and Sparta, 1993). Within wastewater and treatment plants,
volatilisation processes have, in the past, been viewed as a legitimate means for
reducing the total organic load at the discharge point. Moreover, these so-called
'evaporative losses' were incorporated into basic models employed to perform crude
estimates of volatile hydrocarbons emissions from oil/water separator bays
(Litchfield, 1971) and of volatile organohalogens from general wastewaters (Dilling
et al., 1975; Dilling, 1977).
Within the last 10-15 years, the limitations of engineering controls over
volatilisation processes have been recognised as an undesirable feature of public and
industrial sector wastewater management (Dix, 1981). In the U.S.A. uncontrolled
emissions were identified as an ongoing problem in municipal water treatment
plants, with unknown consequences for exposure (Pellizzari, 1982). Similar
problems have arisen in British municipal plants with respect to the reduction of
VOC emissions, particularly concerning odour control. For example, despite the
usage of a variety of 'chemical' controls on malodorous sewage streams (including
scrubbing, oxidation and ozonolysis), the release of volatile compounds, including
sulphur-containing organic thiols, polysulphides and hydrogen sulphide persisted as
a source of concern and complaint (Slater and Harling-Brown, 1986).
Selective improvements in industrial emission control have, however, been made
within the last 10-20 years. In the U.S.A., the petrochemical industry has steadily
developed new technology to reduce VOC emissions. In the early design and use of
primary wastewater treatment bays for industrial oil-water separation (for example
the API Separator; Litchfield, 1971), volatile hydrocarbon emissions arising from
wastewater treatment processes were not subject to control until the use of fixedroof
'vapour-encapsulating' structures and closed-loop vents were mandated by the
USEPA (Vincent, 1979). Within the last decade or so, VOC losses to atmosphere
from large-scale wastewater treatment plants have been the subject of heightened
concern with respect to environmental control and public health, especially in the
U.S.A. Today, the uncontrolled release of VOCs (termed 'secondary fugitive
emissions') are increasingly subject to legislative controls by USEPA and other
regulatory bodies (Springer et al., 1986).
Perhaps one of the most challenging problems encountered by air quality
scientists and occupational hygienists is the lack of suitable, consistent and
scientifically enduring models for estimating emission rates of VOCs from
wastewaters to the local airborne environment. Within the field of environmental
control, appropriate models are needed for estimating transfer of VOCs from water
to airborne compartments as part of emission loss assessments. Reliable models are
also needed for performing residence time calculations on anthropogenically-derived
VOCs as part of ecotoxicological risk assessments (Rogers et al., 1992). Accounting
for VOC losses from water-based processes in the manufacturing industry (for
example paint and dyestuffs) is also required by pollution control legislation.
From an occupational health perspective, predicting volatilisation behaviour
allows evaluation of personal exposure to hazardous VOCs, thus enabling adequate
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Volatilisation processes in wastewater treatment plants 439
controls to be established (for example engineering design, use of respiratory
protective equipment). Inevitably, because programmes of continuous air and water
VOCs monitoring in wastewater plants will be limited in practical terms by
constraints on time and resources, it is clearly desirable to have reliable models for
predicting emission rates of toxic VOCs.
Air/water volatilisation processes
In general, volatilisation models describing air/water interface transfer processes
have not enjoyed widespread use by environmental scientists nor occupational
hygienists in the study of environment and related health effects. Various
commentators have suggested that this may be because such models and the
implied processes are considered too theoretical, mathematically complex or
requiring significant amounts of input data which are not readily available (Ciccioli,
1993; Bianchi, 1994).
As with many predictive models which utilise differing input parameters and
incur major assumptions about natural processes, they may yield variable and often
conflicting results (Wadden and Berrafato-Triemer, 1989). Importantly, the status of
volatilisation models has undergone much change since the early 1990s following
major reappraisals of our understanding of physicochemical and meteorological
parameters which control air-water exchange. In the late 1980s the principal models
in common use to describe volatilisation processes from water bodies were of the (1)
'stagnant film', (2) 'surface renewal' and (3) 'turbulent boundary layer' type. Many
of the commoner so-called 'box' variants are in fact variants of stagnant film and
earlier surface renewal models. Although it is beyond the scope of this paper to go
into these in greater detail, the interested reader can find a summary of each in Liss
and Merlivat (1986).
In particular, the stagnant film model, developed and refined by Broecker and
Peng (1974), was widely applied in the U.S.A. Used throughout the 1970s and 1980s
it also formed the basis of many commercial computer packages for VOC flux
estimating. However, in the early 1990s further development and complementary
fieldwork on surface renewal and turbulent boundary layer volatilisation processes
significantly advanced the validity and effectiveness of models describing
volatilisation processes from water (Upstill-Goddard et al., 1990; Watson et al.,
1991). Simultaneously this new work also highlighted fundamental difficulties in the
stagnant film models. Crucially, the stagnant film model did not accurately predict
the correct dependence of gas transfer on molecular diffusivity nor account for
significant non-linear response in the effects of wind-induced turbulence on
volatilisation rates. The extent to which the model deviated from real situations
varied according to meteorological conditions and the physicochemical properties of
the organic compounds under examination. The literature also indicated that
previous calculations of emission rates based on the stagnant film model may
significantly underestimate actual flux levels, creating major implications for the
reliability of environmental databases founded upon it (Nightingale, 1991).
Some of the key findings of Upstill-Goddard et al. (1990) and Watson et al.
(1991) were later examined and validated through extensive field measurements
(Wanninkhof, 1992). Throughout 1992-93, in one of the first field studies intended
to estimate fluxes of low-level VOCs between environmental (that is estuarine and
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440 A. P. Bianchi and M. S. Varncy
riverine) waters and air under typical environmental conditions, contemporary VOC
emission databases were reprocessed using the new surface renewal models and
found to be up to 10 times higher than predicted by the stagnant film model (Bianchi
and Varney, 1993; Bianchi, 1994).
Given that concentrations of VOCs in wastewater plants are usually much higher
than in marine environments, the consequences for emission rate estimates and their
interpretation in terms of environmental and occupational health are potentially of
greater relevance. Moreover, in this paper we present what we believe to be one of
the first attempts at applying advances in the development of surface renewal and,
where relevant, turbulent boundary layer models to a wastewater treatment bay
under simulated conditions, in the context of an environmental hygiene study.
SURFACE RENEWAL VOLATILISATION PROCESSES—KEY PRINCIPLES
In the study of gas exchange between air and water, the interface between the
two phases is considered as a two-layer (film) system; the main resistance to gas
transport arises from the gas and liquid phase interfacial layers across which
exchanging phases transfer by molecular processes (Liss and Slater, 1974). Since
transfer through the layer system is by molecular diffusion, Fick's first law in the
one-dimensional form (with z as the vertical direction) is applicable, that is
F = —D dc/dz (where F is the flux of gas through the layer, D is the coefficient of
molecular diffusion of gas in the layer material, and c is the gas concentration.
The flux of an organic compound across the water-air interface is a product of
the overall transfer velocity, k, and the extent of the disequilibrium between air and
surface water concentrations (Preston, 1992). The transfer velocity for any given
compound is dependent inter alia on factors such as Henry's Law constant, the
Schmidt number, windspeed and water temperature (Upstill-Goddard et al., 1990;
Watson et al., 1991). Additional parameters which influence the rate of volatilisation
of VOCs across the water-air boundary include aqueous solubility, vapour pressure,
diffusivity, wave action and bubble penetration (Mackay and Yeun, 1983; Liss and
Merlivat, 1986; Wadden and Berrafato-Triemer, 1989).
Some of the most important relationships governing exchange processes can be
summarised as:
F=A:(T)wAC (1)
where AC is the concentration difference driving the flux (F) and &(T)W is the total
transfer velocity (that is the gas exchange constant). The concentration difference is
the difference between the observed aqueous concentration and the calculated
concentration assuming the gas is in equilibrium with the atmosphere and obeys
Henry's law. It can be specifically expressed as:
AC = Ca/T' – Cw (2)
where Ca and Cw are the gas concentrations in air and water, respectively, and H is
the dimensionless and temperature-dependent Henry's Law constant. The total
transfer velocity can therefore be described as:
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Volatibsation processes in wastewater treatment plants 441
where ka and &w are the individual transfer velocities for chemically unreactive gases
in air and water phases, respectively, and a is a factor which quantifies any
enhancement of gas transfer in the water due to chemical reaction. Equation (3) can
be expressed in terms of resistances as:
^(T)w = rvl + r& (4)
where -R(T)w( = 1/^(T)W) is the total resistance, with rw(= \/akw) and ra(= \/Hka) the
resistances of water and air phases respectively.
By substitution of appropriate values for kw, ka and a in Equation (3) it can be
demonstrated that for many gases either rw or ra is dominant. Gases for which rw
is the dominant resistance to transfer mostly have high Henry's Law constants
(that is low solubility) and a is roughly equal to 1.0 [for example CH4, CO2,
CH3I, (CH3)2S)]. This category includes VOCs such as methane and
dimethylsulphide which are often found in industrial and municipal wastewaters,
and for these compounds, kw is the transfer velocity which controls their air/water
exchange.
In this study, most attention was focused on examining the effects of continuous
air movement over the surface of a body of water releasing VOCs to its local
environment using surface renewal concepts, typical of the environment in which
wastewater plants operate. Models of these transfer processes indicate that kw is
proportional to friction velocities in air (£/*) and also to windspeed (£/)• Similarly,
fcw is also proportional to the ratio of the transfer coefficients for momentum
(kinematic viscosity, v) and mass (molecular diffusivity, D) to the power — 2/3 (that
is fcwtfSc"2'3 at U= <5 m s~' or U*= <0.3 m s"1 (that is, a 'smooth surface'
regime) according to the definition of Liss and Merlivat (1986)). Sc is the Schmidt
number, v/D, a dimensionless ratio which is typically in the range of 0.5-2.0 for
gases and 500-2000 for liquids. A useful feature of the Schmidt number is that it
also expresses temperature dependence, notably for liquids in which case Sc
decreases rapidly with increasing temperature, as diffusivity rises and viscosity falls.
To estimate kw it is assumed the surface is smooth and that continuity of stress
across the interface is attained in order to convert the velocity profile in air to an
equivalent profile in water, that is:
kw = 0.082Sc-2/3(ra/rw)1/2£/* (5)
where ra and rw are the densities of air and water. For this model, A:w is proportional
to D ' . Wind tunnel experiments have shown that the relationship between kw and
Sc is not constant, and that under unsteady-state penetration mass-transfer
conditions, a lower dependence is indicated (that is kw is proportional to Sc"1''2)
for a 'rough surface regime' where the actual value of £/«4-13 m s~', and U* «0.3-
0.7 m s " ' (which represents a considerable increase in the slope of kw versus
windspeed). Here, the Schmidt number gives the transfer velocity (that is gas
exchange constant) as proportional to approximately D05. At windspeeds much
above t / = 1 0 m s ~ ' (that is which corresponds to the breaking-wave 'bubble'
regime associated with high winds over the water surface) gas transfer rates are
considerably enhanced, as confirmed by the application of dual-tracer experiments
in rough water environments (Watson et al., 1991).
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442 A. P. Bianchi and M. S. Varney
Assuming the gas exchange constant (k(jyw) is strongly and non-linearly
dependent on the wind speed, three relationships which describe such variation
using carbon dioxide as the model gas, are:
A;(T)W = 0.17(7 (where £/< 3.6 ms"1) (6)
&(T)w = 2.85£/-9.65 (where 3.6 < U 13 ms”1) (8)
where fc(T>w is in cm h”1, and U is in m s~\ This model has been successfully used
to describe air-sea exchange fluxes ranging from low energy water bodies to large
scale open-ocean air-sea exchange associated with high wind speed and significant
surface turbulence (Wallace and Wirick, 1992).
EXPERIMENTAL DETAIL AND MONITORING SURVEYS
Volatilisation measurements were carried out over a 14-month period using a
redundant wastewater holding bay in the eastern Southampton dockland area (that
is an outdoor setting) access to which was made available by a commercial marine
engineering company. Meteorological variables and water temperature were
monitored continuously in addition to the aqueous and airborne concentrations
of selected VOCs, representative of those compounds frequently reported in
municipal and industrial waste streams (Aggazzotti and Predieri, 1986; Lawrence
and Foster, 1987; Thomas et al, 1987; Hazard et al, 1991; Rogers et al, 1992).
Samples were collected over 8-h intervals under a variety of weather conditions
spanning spring through to winter. Under circumstances intended to model severe
water contamination, samples were taken over 15-min periods to estimate the
possible consequences for short-term exposure. Average air temperatures ranged
through the periods of sampling from -21° to 32.2°C.
Experimental methods
The wastewater holding bay was an embedded, fined concrete structure of
approximate dimensions 18×10 m with depth 6.5 m, of which a ‘lip’ 1.4 m high
protruded above a ground-level concrete apron. The bay was gated at either end so
as to allow small vessels (and estuarine water from the Itchen sub-estuary) to enter
at one end and leave at the other. Contaminated water (post-experiment) was
rerouted to a dirty water holding bay. The basic layout shared many similarities in
structure and physical dimensions with municipal or industrial wastewater holdingbays
or oil/water separators. Water depth was recorded during each sampling period
allowing total water volume (and where necessary, residence time) to be calculated.
Water temperature was measured using mercury-in-glass thermometers. Meterological
factors such as relative humidity (%) and air temperature (dry bulb) were
measured using a portable thermohygrograph (Casella Ltd, Bedford, U.K.). Wind
velocity profiles were measured at various locations across the bay using a rotatingvane
anemometer (Casella). Meteorological data were checked daily with the
Southampton Weather Centre. Sampling was avoided during times of prolonged
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Volatilisation processes in wastewater treatment plants 443
rainfall due to the difficulties foreseen in quantitatively accounting for its potential
effects (for example temporary dilution of VOCs in surface water, VOCs ‘washout’
effects of wet precipitation).
Water sampling. Replicate water samples (1 1.) were drawn from the top 0.5 m
depth for analysis at 15-min and hourly intervals for storage in sealed, insulated
boxes cooled to 4°C. The majority of water samples were taken from the mid-point
of the bay (approximately 1 m towards the centre) where natural mixing conditions
were found to be relatively stable and free from internal turbulence (as previously
determined by observation of fluoroscein (BDH-Merck, Eastleigh, U.K.) dye
spiking experiments and water flow measurements using a current meter (Vector
Instruments, Oxford, U.K.). Water samples were taken to the laboratory in sealed
glass vessels (with no headspace) and analysed within 12 h of sampling. Analysis for
VOC content was conducted using a dynamic headspace open-loop multi-sorbent
bed ‘purge-and-trap’ method with automated thermal desorption (Perkin-Elmer
ATD-50, Beaconsfield, U.K.) and BP-1 capillary column-GC with simultaneous
FID/mass-spectral detection. A fuller description of the analytical details and
conditions employed, including calibration procedures, method performance and
quality assurance details, are published in Bianchi et al. (1989), Varney and Bianchi
(1990) and Bianchi et al. (1991) and will not be repeated.
Air sampling. Air samples were collected at a height of 2.0±0.5 m above the water
surface using sampling pumps (low-flow Accuhaler 808 model, MDA, Lincolnshire,
Illinois, U.S.A.; and Flo-Lite pumps, MSA, Pittsburgh, PA, U.S.A.) connected to
Perkin-Elmer ATD-50 sampling tubes packed to the Supelco ‘Carbotrap 300
specification’ [that is Carbotrap C (250 mg) 20/40 mesh; Carbotrap B (175 mg)
20140 mesh; Carbosieve S-III (105 mg) 60180 mesh] supplied by Supelco Inc
(Supelco, Bellefonte, PA, U.S.A.). Air sampling was carried out by suspending
sampling pumps on tripods to which short aluminium ‘poles’ were attached so as to
ensure that the sampling tubes were correctly positioned within the body of air
immediately above the water surface. Air samples were taken to coincide with water
sampling events at 15-min (Flo-Lite pump; pump flow rate = 500 ml min ~’) and 8-hr
intervals (MDA Accuhaler pump; pump flow rate = 50 ml min”1). Air sampling
tubes were capped with Swagelok® end-caps and sealed in glass-jars which were
then stored in separate boxes at 4°C for analysis within 12 h of sampling. Analysis
was carried out on the sampling tubes using Perkin-Elmer ATD-50 thermal
desorption and cGC-FID/MS techniques, very similar to methods used for water
samples. Fuller details of the sampling and analytical methods employed here,
including calibration and quality assurance steps used for airborne VOC samples
were based on the outline protocols given in CONCAWE (1986) and HSE (1989) and
further developed for environmental air sampling by Bianchi and Varney (1993).
Spiking experiments
Sampling measurements were carried out to determine the concentrations of
‘background’ VOCs in estuarine water from which the bay was filled, and the
airspace above it. Aqueous solutions of the VOC compounds of interest were
prepared by dissolving the pure compounds (AnalaR and HPLC spectrophoto-
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444 A. P Bianchi and M. S. Varney
metric grade, Aldrich Ltd, Blandford Forum, U.K.) in water and injecting them into
the bay at a depth of 1 m using a motorised pump and braided steel-mesh hose. For
compounds that were slightly soluble and denser than water, saturated solutions
were prepared, diluted to an appropriate concentration and then injected into the
bay. In further experiments, drums of heavily contaminated water (for example
marine diesel, fuel oil and gasoline-contaminated wastewater) were added in order
to study the variation in volatilisation fluxes arising from the increased levels of
contamination.
RESULTS AND DISCUSSION
A summary of the air and water concentration data for the VOC substances of
interest is shown in Table 1. The data presented in Table 1 are representative of the
‘low concentration’ spiking experiments carried out by adding relatively low-tomoderate
concentrations of each of the VOCs of interest where the concentrated
aqueous solutions added to the bay were typically 5 1. of spike solution
(1000mgl.~’ of each compound), yielding final aqueous concentrations in the
approximate range of 5—15 000 ng l.~’ per component over the experimental
duration (that is JV=73, over 14 months). These concentrations were intended to
represent relatively low levels of contamination found in many typical influent
waters to wastewater bays and their discharge points.
Table 1. Typical air and water concentration ranges of selected VOCs in air (CJ and water (Cw) derived
from wastewater bay spiking experiments
VOC
n-Hexane
n-Dccane
Methylcyclohexane
2,2,4-Tnmethylpentane
Benzene
Toluene
Ethylbenzene
o-Xylene
1,2,4-Tnmethylbenzene
Naphthalene
Carbon tetrachloride
Chloroform
Trichlorofluoromethane
1,1,1 -Trichloroethane
Trichloroethylene
Methyl mercaptan
Dimethyl sulphide
Dimethyl disulphide
Butanone-2
Butanol-2
(Min-Max)*
0.05-33.2
0.07-13.6
0.02-20.2
0.05-24.3
O.KM8.5
0.34-65.7
0.29-60.3
0.36-58.3
0.40-50.0
0.44-28.3
0.23-27 3
0.77-99 3
0.01-20 2
0.75-^7.5
1.21-59.4
0.06-95.3
0.05-99.2
0.10-93.3
0.35-8.5
0.28-7.4
c.
(Mean)f
12.3
8.4
7.6
11.0
25.3
33.7
30.3
28.4
26.7
13.4
15.6
47.0
9.3
25.3
30.3
44.5
55.3
50.3
4.5
3.5
GMJ
6.0
3.3
3.0
5.6
11.9
15.9
14.1
13.3
12.3
6.5
7.3
23.5
4.5
12.2
14.3
20.4
27.5
25.6
2.3
1.8
(Min-Max)*
8.8^400
2.7-2203
3.3-1920
2.1-2504
10 4-8320
17.4-10122
15.2-10001
12.8-11030
10.2-8684
5.3-3686
12 3-6478
93 4-21300
15 6-3647
10.2-8500
20 3-9403
27 3-11 020
18.2-16304
29 3-13000
15.6-6900
18.8-7530
Cw
(Mean)f
1020
964
873
1285
4672
5602
4975
5204
4664
1405
3692
10203
1502
4633
4304
6422
8377
7002
3500
3790
GMJ
945
467
420
640
2202
2829
2704
2474
2645
693
1578
4365
745
2403
2102
3404
4355
3680
1200
1540
The data shown are based on a pool of A’= 73 ‘low concentration’ experiments and relate to 8-h timeweighted
average concentrations. Concentration data are expressed as ng l.~’
‘Arithmetic range shown as Min-Max values obtained.
tAnthmetic mean
JGeometnc mean.
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Volatilisation processes in wastewater treatment plants 445
The highest recorded concentration was for chloroform (Ca = 99.3 ng l.~’,
corresponding to CW = 213OO ng l.~’), a major contribution of which came from the
influent estuarine water used to flood the bay. This is consistent with the results of
Aggazzotti and Predieri (1986) who identified comparable levels of chloroform in
municipal water streams where it also represented the most ubiquitous compound
found in the highest concentration. Elevated levels of chloroform had previously
been reported by Bianchi (1994) and Dawes and Waldock (1994) in the Itchen subestuary
to Southampton Water, mainly as a by-product of water chlorination and
industrial processes. Organosulphide levels in ‘background’ water (and in air
samples) were also higher than anticipated. Up to 25% of the total amount of
methyl mercaptan, dimethylsulphide (DMS) and dimethyldisulphide (DMDS)
measured in the bay were also derived from surface water in the Itchen. Water
quality varies significantly within the estuary (Soulsby et al., 1985), highlighting the
need to account for extranneous sources of VOCs. The sources of the
organosulphides were traced to a nearby sewage outfall and unrelated biological
decay processes of plankton such as the photosynthetic ciliate {Mesodinium rubrum).
Among the alkylbenzenes, for example, toluene was also found in the highest
concentrations, particularly in air. The lowest airborne concentrations recorded
were for butanone-2 and butanol-2, which may be expected given their hydrophilic,
polar nature.
Theoretical equilibrium concentrations
Given the analytical data set made available by the air and water concentration
surveys, it was possible to estimate concentration differences (AC) across the airwater
interface. By using the theoretical Henry’s Law constants for each compound
(Equation 2), theoretical equilibrium concentrations were calculated (that is the
aqueous concentration in theoretical equilibrium with the concentration measured
in air) using the Henry’s constant protocol described by Nightingale (1991). A
summary of the theoretical equilibrium concentration values are shown in Table 2.
For all compounds studied, the actual aqueous concentration significantly exceeded
the theoretical air-water equilibrium concentration predicted by achievement of
interphase equilibrium, showing that, despite the relatively low range of aqueous
concentrations, the water was supersaturated with respect to atmospheric transfer.
Thus, in terms of Equation (2), Cw is mainly >>CaH~l. At these aqueous
concentrations, VOC movement would therefore be highly unidirectional (that is
from water to air), with the atmospheric ‘compartment’ representing the major sink.
This is an important finding, since under different environmental conditions the
direction of transfer may be reversed (that is from air to water). For example it was
previously shown by Bianchi and Varney (1993) that under certain conditions
following episodes of elevated airborne pollution, vapour-phase toluene may cross
the ‘air-water’ interface, representing a source of toluene to surface waters (that is
where CJJ~’ > Cw.
VOC flux calculations
Fluxes were derived by calculating the respective Co and Cw data for each
component, allowing for maximum and minimum recorded wind speed across the
surface of the water body. Using this data, gas exchange constants kmv/) were
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446 A. P Bianchi and M. S. Vamey
Table 2. Equilibrium concentration values for representative air and water VOC concentration data
voc
n-Hexane
n-Decane
Methylcyclohexane
2,2,4-Trimethylpentane
Benzene
Toluene
Ethylbenzene
o-Xylene
1,2,4-Trimethylbenzene
Naphthalene
Carbon tetrachloride
Chloroform
Trichlorofluoromethane
1,1,1 -Trichloroethane
Trichloroethylenc
Methyl mercaptan
Dimethylsulphide
DimethyWisulphide
Butanone-2
Butanol-2
Estimated equilibrium
concentration
0.26
0.03
0.66
0.12
210.8
187.2
121.2
118.3
133.5
1595.2
14.4
522.2
1.86
105.4
101.0
111.3
1843.3
251.5
225.0
3500.0
Measured aqueous
concentration*
1020
964
873
1285
4672
5602
4975
5204
4664
1405
3692
10232
1502
4633
4304
6422
8377
7002
350
379
Calculated Henry’s
Law constantf
47
252
11.4
91.2
0.12
0.18
0.25
0.24
0.20
0.0084
1.08
0.09
5
0.24
0.3
0.4
0.3
0.2
0.02
0.01
Values are estimated from arithmetic mean data for air ( O and water (Cw) for the data set ( # = 73).
Concentration data are expressed as ng I.”1.
‘Measured aqueous concentration ng I.”1.
fHenry’s Law constants calculated according to Liss and Slater (1974), Sauer (1978) and Nightingale
(1991) corrected for temperature and pressure.
calculated for each component under the prevailing conditions. As these values
relate to the model gas CO2, a correction must be applied for other volatile
compounds to compensate for differences in molecular diffusivities. There are two
ways of accomplishing this. Firstly, by using the ratio of the square roots of the
molecular weights of CO2 and the organic compound, as detailed by Liss and Slater
(1974) and Nightingale (1991); or secondly, by examining the variation of the gas
exchange constant with both the compound of interest and temperature, which can
be described by a power dependence on the Schmidt number as Ki/K2= (Sc^ScO”
(Watson et al. (1991). For a windspeed of more than 3.6 m s ~ ‘ , the power
dependence (n) is approximately 1/2. As there are few predetermined values
available for molecular diffusivities, the molecular mass correction was applied to
values of A^-^ in this body of work.
Flux transfer rates were significantly a function of wind speed and VOC
concentration differences between air and water phases. In the Southampton Dock
area, wind direction is usually from the south-west for more than 75% of the year,
and normally exceeds 2.3 m s~’ for more than 80% of the time (Bianchi, 1994).
During winter months (that is which we classed as December through to March),
wind speeds during sampling exercises ranged from 3.6-13.9 m s~’. In summer
months (that is classed as May to September) wind speeds were much lower, from
1.9-4.8 m s”1.
Within the ranges of concentration tabulated in Table 1, fluxes for a broad range
of VOCs ranged from approximately 0.04xl0~8 to 9.0xl0~8 g cm”2 h~~’. A range
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Volatilisation processes in wastewater treatment plants 447
Table 3. Flux estimates for typical winter (5/2/95) and summer (23/7/95) weather conditions
voc
n-Hexane
n-Decane
Methylcyclohexane
2,2,4-Trimethylpentane
Benzene
Toluene
Ethylbenzene
o-Xylene
1,2,4-Trimethylbenzene
Naphthalene
Carbon tetrachlonde
Chloroform
Trichlorofluoromethane
1,1,1 -Trichloroethane
TricbJoroethylene
Methyl mercaptan
Dimethybulphide
DimethyWisulphide
Butanone-2
Butanol-2
c.
18.9
11.4
17.1
20.1
41.9
60.7
58.3
53.7
45.6
20.3
15.1
80.0
15.1
36.9
445
58.3
69.7
67.4
4.2
4.9
(5/2/95)’
Qv
3295
2105
1763
2209
7095
8654
7920
7562
6827
3057
4929
15300
3005
6250
8133
8745
14321
11450
5880
5956
10.21
7.95
9.58
8.87
10.34
9.87
8.79
9.19
8.64
8.36
7.64
8.63
8.08
8.20
8.27
13.66
12.03
9.76
11.16
11.01
FluxJ
336.4
167.3
168.8
195.9
692.8
820.9
675.7
674.4
570.2
76.2
375.4
1243.8
242.6
499.9
660.4
1181.5
1694.9
1084.6
632.8
116.3
c.
9.8
4.7
5.0
6.1
15.0
22.3
23.5
20 4
19.3
111
120
39.3
8.1
20.0
24 3
30.1
50.2
42.3
2.1
1.8
(23/7/95)t
560
420
440
620
2150
2720
2000
1950
1920
1700
1495
5500
655
2350
2090
3605
4107
3595
1190
1904
K 0.3 m s”1)
and U* = (6.1 + O.63t/iO)o5t/,o.
(where U\Q is the wind speed at a nominal height of 10 m).
Benzene, for example has a ScL= 1021 (at 20cC). By employing a typical value of
0.42 (of [/*) for winter, a flux rate of 16.5 cm h”1 is obtained. The ScL value
increases with increasing alkylation of the benzene ring (for example ScL
toluene = 1155) and hence flux rates are progressively lower for C3- and C4-
alkylbenzenes. ScL increases with decreasing water temperature and hence broad
estimates can be made of the variation of flux rate with water temperature. The
precise quantitative relationship between Schmidt number, temperature and related
environmental variables is not yet fully understood but it can be shown, for example
that if ScL increases by 2.8% per degree Centigrade decrease, the Schmidt number
for benzene at typical winter temperatures of about 70°C is predicted to be about
1650. At this theoretical value, the magnitude of Ktjy* (benzene) decreases from
16.5 cm h”1 (at 20°C) to 13.0 cm h”1.
Individual flux rates are also highly dependent on the absolute concentration of
contaminant in the water phase. This may be illustrated by comparing net fluxes
from water bodies which contain moderate concentrations of VOCs with those from
relatively pristine water bodies such as open sea water. For example, Nightingale
(1991) performed identical estimates for volatile organohalogens in the southern
North Sea, and estimated a net daily flux of 0.1 lx 10~10 g cm”2 h~’ for
chloroform. By way of comparison, the net daily fluxes obtained for chloroform
in our study were up to 1244xlO~10 g cm”2 h~\ between 150 and 12000 times
higher for summer and winter months respectively, highlighting the significant
differences in rate and total mass of transfer which is theoretically possible for the
same compound.
In terms of occupational health and environmental issues, these results also
suggest that volatilisation processes within wastewater bays may, under conditions
determined by AC=Ca//~’-Cw, their respective concentrations, ^mw, and air and
water temperatures, represent important factors which may determine personal
exposure to individual (or total) VOCs. For example, crude estimates based on the
experimental data from this study suggest that where ECW>5 mg I.”1, (as total
VOCs) in water (that is levels commonly encountered in wastewater plants), £Ca
may reasonably be expected to reach or exceed 50 mg m~3 (as total VOCs) in air
within the activity zone of workers, under moderate wind speed conditions (for
example 3-6 m s~’). More precise estimates of exposure to individual compounds
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Volatilisation processes in wastewater treatment plants 449
should be readily available when performing finer estimates. In theory, it should also
be possible for occupational hygiene personnel to make sufficiently useful
predictions of likely airborne concentrations based on measurements (or good
estimates) of aqueous VOC concentration and windspeed alone.
Contamination (spill) events
Importantly, increased loadings of contaminant material (including hydrocarbon
or solvent-based fluids) are relatively common occurrences in municipal and
industrial wastewater plants and can take place at any time, particularly after a spill
event. Such circumstances may arise after a process plant upset or following a civil,
traffic-related or industrial incident (for example fire, emergency release). Depending
on the nature of the VOCs in the influent water and prevailing weather conditions,
these factors will have a major impact on volatilisation rates. Further experiments
were carried out by injecting large volumes (200-300 1.) of contaminated effluent
water from local shipping activities (for example marine fuel oil, gasoline and dieselcontaining
ballast water) into the bay from large steel drums using drum pumps.
Sampling was carried out within 15-min time ‘windows’ within 5 min of charging
the bay, and for longer periods of time extending up to 8 h. Representative results of
a sample ‘injection’ of a mixture of ballast water severely contaminated by kerosine
and raw gasoline are shown in Table 4.
The results indicate that at relatively moderate values of K^-^ (that is 2.3-
3.7 cm h~’) proportionally high flux rates can be achieved ‘driven’ mainly to the
significant gradient (high AC) between air and water-phase concentrations (in
AC=CaH~i — Cw), especially in the first 15 min. Moreover, as the results suggest,
personnel working within the immediate zone would probably require respiratory
protection or limited work periods to prevent or reduce the risk of high single- and
Table 4. Flux estimates for a simulated contamination event following discharge of wastewater contaminated
by petroleum-fluids
VOC
n-Hexane
n-Decane
Benzene
Toluene
Ethylbenzene
o-Xylene
1,2,4-Trimethylbenzene
Naphthalene
c.t
111
52
156
242
219
215
127
7
(15-min
cw§
124
183
137
106
105
104
84
48
period)*
%)«
2.86
2.26
2.91
2 78
2.50
2.58
2.42
2.36
Fluxi
35
41
40
43
26
27
26
11
c.t
2.0
0.9
3.9
4.7
3.5
4.1
0.9
0.1
(8-h
cw§
3.5
4.7
5.1
6.7
1.5
2.1
0.7
0.3
period)t
^(T)w
3.67
2.90
3.72
3.57
3.21
3.31
3.10
3.03
Fluxi
1.3
1.4
1.9
2.4
0.5
0.7
0.5
0.1
Air and water concentration ranges of principal VOC are shown for air (C.) and water (Cw)
concentrations derived from sampling and analysis. Experimental data are based on 15-min and 8-h
sampling periods. K<x)w are expressed a s c m h " 1 corrected according to mean wind speed. Concentration
data are expressed as ng 1."'.
*Mean wind speed (15-min period) = 4.8 m s ~ \ mean air temp 15-min= 17.5°C, mean water
temperature = 10.2°C.
fMean wind speed (8-h) = 5.2 m s~', mean air temp (8-h) = 18.2°C, mean water temperature = 10.7°C.
JC« expressed as x 103 ng I."1.
§CW expressed as xlO6 ngl."1 .
> CaH~l). For all alkylbenzenes, values of
^(T)w reduce with increasing alkylation of the benzene ring. Given the relatively
greater chronic health risks associated with benzene compared to its alkylbenzene
counterparts, this feature would need to be adequately addressed during exposure
risk assessment.
Among the volatile organosulphides, values of K<j)w are highest for most low
molecular compounds with high vapour pressures (for example methyl mercaptan,
DMS). In particular, these factors may help to explain the labile nature of highly
malodorous methyl mercaptan (including ethyl and propyl mercaptans) within
municipal sewage works and waste transfer and recycling stations where these
compounds are frequently found, enhancing the probability of public odour
nuisance complaints (Slater and Harling-Brown, 1986; Bianchi, 1994).
Taken together, the foregoing discussion points suggest that a valid exposure
risk assessment would require the occupational hygienist to obtain a sound database
of aqueous concentration data in addition to airborne data, the latter activity being
the 'normal' premise of the practising hygienist. From our experience in assessing
volatilisation behaviour in open waters, wastewater lagoons, waste treatment plants
and industrial oil-water separators, it seems a valid and achievable precaution to
develop local predictive models which describe volatilisation fluxes and anticipated
airborne concentrations of commonly encountered VOCs under a range of
operating conditions regarded as 'routine' for the plant in question. Furthermore,
these results reinforce the necessity in considering the use of controlled
encapsulation, venting and filtration systems as a management strategy for VOCs
in wastewaters, as opposed to a dilute-and-disperse approach which calculations
indicate may exacerbate an already existing inhalation risk or enhance a nuisance
odour risk for certain types of VOCs.
Research carried out as part of this and earlier studies indicates that these
volatilisation models should also usefully apply to indoor water bays, tanks or
receptacles as much as they do to outdoor water bodies. New work on indoor paint
spraying tasks (unpublished at the time of writing) indicates that the surface renewal
model predicts airborne concentrations of VOCs (at levels of approximately 20-
Downloaded from http://annhyg.oxfordjournals.org/ by guest on May 11, 2012
Volatilisation processes in wastewater treatment plants 451
500 mg m~3 +25%) associated with 'wet collection' drainage sluices used in
manual spraying activities adjacent to spray booths.
Perhaps one of the main factors which detracts from the validity of the model
lies in the existence of competing 'sinks' for VOCs. For example, the presence of
high concentrations of waterborne suspended solids in treatment plant bays
(inorganic flocculant, sand particles, organic sewage particles) may act as selective
adsorbent sites for hydrophobic VOCs, reducing the solute fugacity or partial
pressure, and so reducing the volatilisation rate. Adsorption to sediments results in
removal from the liquid phase and incorporation into a solid phase. Exchange may
also take place with and between resuspended particles at varying rates, depending
upon a variety of parameters. However, in a series of field measurements and
modelling experiments examining the extent of adsorption within water treatment
plant effluents, Bianchi (1994) demonstrated that even under conditions of high
suspended solids (that is 1200 mg I."1) with correspondingly high fractional organic
contents (that is approaching 100%), the maximum amount of VOCs removed by
adsorption was 10% (of total mass) for chloroform, and 5% (of total mass) for most
volatile aromatics (for example including benzene, toluene and C2-alkylbenzenes)
and organohalogens (for example including chloroform and carbon tetrachloride),
where actual aqueous concentrations were between 1 and 10 //g I."1. It is therefore
unlikely that adsorption would represent an important competing sink nor
significantly reduce volatilisation rates under conditions in which inhalation
exposure represented a potential risk.
Photo-oxidation processes have also been considered as a potential 'sink' for
VOCs. It is, however, unlikely that photo-oxidation reactions would occur at fast
enough rates to remove VOCs. Many of the final photo-oxidised products of
anthropogenic volatile hydrocarbons are aldehydes (Grosjean et al., 1978; Howard
et al., 1991). In particular some alkanes and alkenes oxidise to saturated aldehydes
(Cox et al., 1980), none of which were identified during this study.
SUMMARY AND CONCLUSIONS
This study has provided new and quantitative experimental data describing the
variation in gas exchange constants and water-to-air fluxes for a broad range of
hazardous VOCs commonly encountered in municipal and industrial wastewaters
under varying environmental conditions. The data were derived using a relatively
new volatilisation model which representatively accounts for the effect of
physicochemical characteristics and meteorological factors exerting major controlling
influences on volatilisation behaviour. With hindsight, the value of obtaining a
broad experimental data set over a year under changing climate conditions and in an
environment which closely resembled a real wastewater bay (as opposed to
laboratory simulations) cannot be underestimated. We believe that, despite potential
gaps in the experimental protocol, the data provide a useful platform from which to
examine the relationship between gas transfer velocities and wind speed as predicted
by Upstill-Goddard et al. (1990), Watson et al. (1991) and Wanninkhof (1992).
Further confirmation or comparison of the experimental findings was made difficult
by the apparent lack of data in the literature concerning similar studies of this
Downloaded from http://annhyg.oxfordjournals.org/ by guest on May 11, 2012
452 A. P. Bianchi and M. S. Varney
nature, which may indicate that this topic remains a new and still emerging field in
the discipline of occupational hygiene.
For all compounds we studied, the aqueous concentrations used to model typical
wastewater bay conditions significantly exceeded the theoretical air-water
equilibrium concentration predicted by achievement of interphase equilibrium.
Despite what we considered to be quite low absolute aqueous concentrations of
VOCs, it seems likely that most wastewaters will be supersaturated with respect to
atmospheric transfer. In most cases, aqueous concentrations of VOCs in wastewater
bays would probably exceed their theoretical air-water equilibrium concentrations.
The total set of VOCs data, including concentration differences (AC) between air
(Ca) and water (Cw) phases, gas exchange constants (kmv/), and fluxes were
calculated using the most up-to-date 'surface renewal' equations. To the best of our
knowledge, these models have been proven to be reliable although new work is
needed to examine their applicability to occupational environments and in particular
to a wide range of waterborne compounds used by industry. Within the ranges of
concentration commonly encountered in this study, fluxes for most VOCs ranged
from approximately 0.04 x 10~8 to 9.0×10~8 gem"2 h~', the highest values usually
being observed with alkylbenzene and organosulphide compounds. Furthermore,
given the higher concentrations of VOCs in wastewaters during winter associated
with greater windspeeds immediately above the air-water interphase, values of Kmw
were 20—30 times higher than in summer. These findings reinforce general
conclusions that programmes designed to minimise VOC emissions to air must
account for mechanisms which enhance volatilisation and incorporate precautionary
strategies around organic vapour control.
The results of our basic experiments intended to represent 'contamination spill
events' indicated that at relatively moderate values of Kfjy,, (that is 2.3-3.7 cm h~')
high flux rates are observed due to the significant gradient (high AC) between air
and water-phase concentrations (that is where AC=CaH~1 — Cw), especially in the
first 15min. Moreover, as our results highlight, personnel working within the
immediate vicinity may be at risk of higher levels of exposure than anticipated
during such occurrences.
We believe that these data should serve as a useful platform from which to
encourage the use of volatilisation modelling by occupational hygienists carrying
out risk assessments in a variety of wastewater treatment operations. Through
appropriate application of the theoretical models described, reasonable predictions
and estimates can be made of potential VOC exposure scenarios to workplace
personnel employed in wastewater storage and treatment activities.
Acknowledgements—The authors would like to thank Mr J. Leach and Mr A. Chidwick (Condive Marine
Co., Southampton) for access to and use of the water bay utilised for the field experiments, for assistance
in operating the equipment and arranging supply of the contaminated wastewater materials. The authors
would also like to thank Prof. P. Liss and Dr P Nightingale (School of Environmental Sciences,
University of East Anglia) for helpful comments on the use of the volatilisation models.
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Sanitation Systems: Pros and Cons

The UV radiation damages the cell’s DNA, thus rendering the bacteria incapable of reproduction and the bacteria dies off.

Thermal disinfection

Functional principle:
The complete system including the withdrawal fittings are heated up to > 70 °C for a period of at least 3 minutes.

Advantages:

  • Provided the heating capacity is available, this can be done at any time

Disadvantages:

  • Risk of scalding
  • Increasing scale precipitation
  • Scale precipitation generates the medium for legionellae
  • Risk of corrosion in the piping system
  • High organisational and personnel effort required

Chemical disinfection

Functional principle:
Dosing of high chlorine concentrations, at least 10 mg/l free chlorine. The DVGW work sheet W 291 must be observed.

Advantages:

  • Feasible from the technical point of view as long as the facility has been shut down or is in disuse.

Disadvantages:

  • High demand regarding personnel and chemicals
  • Risk of corrosion in the piping system
  • Problematic disposal
  • Legionellae located in host organisms are not eliminated completely.
  • A minimum concentration of 50 ¿ 100 mg/l of chlorine is required to destroy the host organisms.

Chemical disinfection – Chlorine electrolysis

Functional principle:
Electrolytic generation of hypochlorous acid (free chlorine), distribution of the chlorine by means of existing circulation pipes.

Advantages:

  • Operation even at lower temperatures is possible

Disadvantages:

  • The electrolytically generated values of max. 0,3 mg/l of free chlorine are not adequate to kill legionellae
  • If not enough chloride is contained in the water, additional dosing of chloride is required.
  • Due to the electrolytic processes hydrogen is generated which in turn leads to corrosion.

 

UV irradiation

Functional principle:
Irradiation of the water by means of UV-C light at a wave length of 254 nm

Advantages:

  • Immediate destruction of legionellae that are floating free in the water

Disadvantages:

  • Legionellae located in protective shelters (amoebae) are not destroyed

Ultrasound/UV treatment

Functional principle:
Combination of ultrasound and UV. The ultrasound is used to break up the amoebae thus setting free the legionellae which will then be killed by means of UV light.

Advantages:

  • Unit may be installed in new as well as in existing systems, easy system integration
  • Operation even at lower temperatures is possible
  • No scale precipitation and corrosion problems
  • No dosing of chemicals
  • Legionellae located in protective shelters (amoebae) are killed as well

Disadvantages:

  • Sizeable space required