Ozonation of Municipal Wastewater Effluents

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STATE-OF-THE-ART REVIEW

Ozonation of Municipal WastewaterEffluents

Panagiota Paraskeva, Nigel J. D. Graham

ABSTRACT: The increasing use of ozone in the treatment of

municipal wastewater effluents has been stimulated by the need toachieve higher effluent quality and greater compliance with

physicochemical and microbiological quality standards before discharge.

These standards are applied when the effluent may pose a risk to the

public through direct contact and where the effluent is used for

agricultural purposes or water reclamation. Although various alternative

technologies exist for upgrading wastewater effluents, ozone treatment

may be the most appropriate approach in particular cases. This review

summarizes the current status of the use of ozone for treating municipal

effluents with respect to disinfection efficiency, its effect on the

treatability of the effluent and on aggregate effluent parameters, the

potential for the formation of ozonation byproducts, and its effect on the

toxicity and mutagenicity of the effluent. The importance of treatment

conditions (e.g., contact time) is also reviewed. Water Environ. Res., 74,

569 (2002).

KEYWORDS: wastewater, effluent, municipal, treatment, disinfection,

ozone, oxidation, byproducts, toxicity.

Introduction

Ozone (O3

) is a highly reactive chemical with a high oxidation

reduction potential (E0

H) of 2.07 V (Hoigne, 1998). Its use in

aqueous conditions usually leads to the simultaneous production of

secondary oxidants, such as radical species (OH), whose oxida-

tion power is much greater than molecular ozone (Hoigne and

Bader, 1985). This factor, taken with the absence of any halogen

constituent, has made ozone a valuable chemical in water and

wastewater treatment. The primary field of ozone application is the

water industry, although ozone is increasingly being applied to

solving gaseous and liquid problems in other industries (e.g.,

textiles, paper and pulp, and cooling waters), and in medical

therapy. In water treatment, ozone can be used in various steps of

the treatment process to enhance biological processes and micro-

flocculation, iron and manganese removal, degradation of pesti-

cides and other micropollutants, and taste and odor removal.

Ozone is particularly beneficial for the disinfection of potable

water because it is effective against both bacteria and viruses and

can also remove cysts and eggs (Anderson, 1997; Langlais et al.,

1991; Rice and Netzer, 1984); currently, there is considerable

interest in its effectiveness against the Cryptosporidium parvum

cyst, in particular. Camel and Bermond (1998) recently completed

an extensive review of the use of ozone in drinking water.

In the context of wastewater treatment, the high reactivity of

ozone makes it appropriate for achieving certain objectives whenapplied either alone or in combination with other processes (e.g.,

filtration). These objectives relate to either the need to achieve

higher quality standards prior to final discharge or to meeting

standards for effluent recycling. Specifically, the objectives may

include color removal, disinfection, the degradation of organic

micropollutants, the conversion of hard chemical oxygen de-

mand (COD), and effluent oxygenation. The heterogenous nature

of municipal wastewaters and the economic limitation of ozone

application make it unlikely that organic substrates can be com-

pletely degraded (to carbon dioxide and water) by ozone treatment.

This has led to concern over the presence of intermediate byprod-

uct compounds that may be of toxicological significance. Many

studies have been conducted concerning the reactivity of ozoneand its secondary oxidants with a wide range of inorganic and

organic substances. Details of these studies, including rate con-

stants, can be found in recent reviews in the literature (Hoigne,

1998).

Similarly, the reactivity of ozone with humic substances has

received considerable attention in recent years because such sub-

stances are found in all kinds of natural and polluted waters and are

known to influence ozone decomposition and the presence of

secondary radicals. Ozone causes substantial structural changes to

humic substances such as a strong and rapid decrease in color and

UV-absorbance resulting from a loss of aromaticity and depoly-

merization; a small reduction in total organic carbon (TOC) (e.g.,

10% at 1 mg O3/mg C); a slight decrease in the high apparentmolecular weight fractions and a slight increase in the smaller

fractions; a significant increase of the carboxylic functions; and the

formation of ozonation byproducts. These byproducts are mainly

aldehydes (formaldehyde, acetaldehyde, glyoxal, and methylg-

lyoxal) and carboxylic acids (formic, acetic, glyoxylic, pyruvic,

and ketomalonic acids) (Camel and Bermond, 1998). In addition,

the results available to date indicate that the ozonation of humic

substances tends to increase their biodegradability, although di-

rectly quantifying the extent of this increase is difficult because of

the variety of methods used to measure biodegradability. Some of

these studies have been described by Langlais et al. (1991) and

others.

This review summarizes the current status of the use of ozone

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for treating municipal effluents with respect to disinfection effi-

ciency, its effect on the treatability of the effluent and on aggregate

effluent parameters, the potential for the formation of ozonation

byproducts, the effect of ozone on the toxicity and mutagenicity ofthe effluent, and the influence of the ozone-contacting conditions

on performance.

Disinfection of Municipal Effluents

In municipal wastewater treatment, ozone has been used pri-

marily for effluent disinfection. Generally, the need for effluent

disinfection has arisen through the enactment of legislation that

has defined specific microbiological standards for effluent dis-

charge quality. These standards not only vary with the specific

country or region of application, but they also depend on whether

the effluent is to be discharged to a sensitive waterbody or used for

agricultural or reuse purposes. Within countries of the European

Union (EU), current legislation concerning effluent discharge to

receiving waters involves the EU bathing waters directive (CEC,

1976), which sets microbiological and physicochemical standards

for inland and coastal waters designated as bathing areas (Table 1).

Individual effluent limits are, therefore, set by local regulatory

bodies to comply with these standards. In the United States, the

California requirements for total coliforms are 1000/100 mL (80th

percentile) for coastal bathing waters (equivalent to a median of

230/100 mL), 70/100 mL (median) for shellfish growing areas, and

23/100 mL (median) for confined waters used for bathing or

recreation (White, 1999). A common effluent discharge standard in

the United States is 200 fecal coliforms/100 mL (Rice, 1997). For

effluent reuse, the World Health Organization (WHO) guidelines

on effluent quality (WHO, 1989) define a maximum value of 1000

fecal coliforms/100 mL of effluent (on a geometric mean basis)

and less than 1 viable intestinal nematode egg/L (on an arithmetic

mean basis) for unrestricted agricultural reuse. In the United

States, specific criteria for effluent quality for wastewater recla-

mation and reuse are set by individual states. However, the Cali-

fornia Title 22 regulations (State of California, 1978) are recog-

nized as being particularly comprehensive and, as such, have been

adopted by some European countries (Liberti and Nortanicola,

1999). The California regulations require physicochemical treat-ment and disinfection to achieve a limit of 2.2 total coliforms/100

mL (median) and a 5-log removal of virus.

Although ozone is approved for short-term field trials only in the

United Kingdom, in other countries ozone has been applied to

wastewater effluents primarily for disinfection. Ozonation has also

been used for COD polishing (i.e., lowering of residual concen-

trations) and enrichment of effluent with oxygen. In the United

States, the first ozonation plant for the disinfection of municipal

effluent was built in 1975. Since then, at least 45 ozone plants have

been built. However, in recent years the number of plants has

declined because legislation has changed and, in most cases,

disinfection is no longer required (Robson and Rice, 1991). There

are two operational ozone plants in Canada (Larocque, 1999) andseveral in Korea. There are approximately 200 ozonation units in

Japan treating wastewater effluent and approximately 400 units

treating night soil wastewater effluent (Matsumoto and Wa-

tanabe, 1999). In Europe, there are two major ozone plants in

France and 134 plants in Germany treating municipal wastewater

effluent and exhaust air. The use of ozonation for the disinfection

of effluents before agricultural use is also increasing in southern

Europe (Bohme, 1999; Le Pauloue and Langlais, 1999; Liberti et

al., 1999). In the United Kingdom, there is currently one waste-

water treatment plant treating its secondary effluent with ozone for

color removal (Churchley and Upton, 1997).

Ozone has been found to be very effective at inactivating a wide

range of microorganisms. The mechanism of bacterial inactivation

by ozone is thought to occur by general inactivation of the whole

cell. Thus, ozone causes damage to the cell membrane, to the

nucleic acids, and to certain enzymes. Ozone is particularly effec-

tive against viruses, where its use can achieve the highest standards

(Tyrrell et al., 1995). The mechanism of viral inactivation involves

coagulation of the protein and oxidation of the nucleobases form-

ing the nucleic acid. Protozoan cysts, specifically Giardia and

Cryptosporidium, and bacterial spores are more resistant to ozone

than bacteria and viruses, although moderate degrees of inactiva-

tion (Table 2) have been demonstrated under realistic ozonation

conditions (Owens et al., 2000). It has been reported that micro-

organism reactivation after ozonation is unlikely to occur (CES,

1988; U.S. EPA, 1986), and Lazarova et al. (1998) did not observe

any measurable regrowth of Escherichia coli in seawater after

Table 1Microbiological standards for designated

bathing areas in the European Union (CEC, 1976).

Microorganism

Guideline value

(G) (80th

percentile)

Mandatory value

(I) (95th

percentile)

Total coliforms

(count/100 mL) 500 10 000Fecal coliforms

(count/100 mL) 100 2000

Fecal streptococci

(count/100 mL) 100

Salmonella (count/L) 0

Enteroviruses (plaque

forming units/10 L) 0

Table 2Inactivation of microorganisms by pilot-scale ozonation (Owens et al., 2000).

Microorganism

Temperature

(C) pH

CTa

(mgmin/L)

Log inactivation

range

Bacillus subtilis endospores 22.7 1.0 7.93 0.32 0.7018.35 02.17

Cryptosporidium parvum oocysts 24.5 1.6 8.24 0.20 2.557.15 0.572.67

Cryptosporidium muris oocysts 23.6 1.6 8.40 0.11 0.9810.7 0.362.56

Giardia muris oocysts 25.2 1.1 7.57 0.29 0.281.04 1.522.70

Poliovirus 1 25.0 1.0 8.05 0.17 0.192.49 1.433.85

a Concentration time (CT) product, based on integrated dissolved ozone concentration values (C) and theoretical residence time (t).

Paraskeva and Graham

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ozonation. A more complete summary of the specific mechanisms

of microbial inactivation can be found in Langlais et al. (1991).

As a disinfection method for wastewater effluents, ozone is

applied mainly to secondary-or tertiary-treated effluents because

the ozone demand of primary or raw effluents is so large that it

makes the use of ozone not cost- and energy-effective. The effec-

tiveness of disinfection depends on the ozone dose, the quality of

the effluent, the ozone demand, and the transfer efficiency of the

ozone system. An example of the dependence of disinfection on

ozone dose and the nature of the effluent is shown in Figure 1. The

disinfection dose (i.e., the dose of ozone that achieves certain

microbiological standards in a municipal effluent) is expressed as

the transferred (or absorbed) mass of ozone per liter of effluent in

milligrams per liter. Another form of characterization for disinfec-

tion conditions is the C t product, where C is the concentrationof dissolved (residual) ozone measured at the outlet of the contact

chamber (in milligrams per liter) and t is the contact time between

the residual of ozone and water (in minutes). However, this char-

acterization may not be appropriate for effluent disinfection be-

cause significant bacterial inactivation can be achieved (e.g., 2 or

3 logs) prior to an ozone residual appearing (Absi et al., 1993;

Janex et al., 2000; Paraskeva and Graham, 1999). The physico-

chemical quality of the effluent is particularly influential in deter-

mining the effectiveness of disinfection and the ozone dose re-

quired to achieve a specified performance. The organic content of

the effluent, as expressed by COD and total suspended solids

(TSS), is commonly used as an indicator of the ozone dose

required for disinfection of secondary effluents (unfiltered) and the

likely overall performance (Table 3). Attempts have been made to

establish empirical relationships or formulas to predict the total or

fecal coliform inactivation by ozonation in terms of the organic

and inorganic species, such as COD, TSS, and nitrite-nitrogen

(NO2

-N). Absi et al. (1993) found a close linear relationship

(correlation coefficient 0.95) between the logarithm of fecalcoliform survival (counts remaining/initial counts) and the influent

COD for a chemically assisted primary wastewater, although this

was for a very narrow ozone dose range (8 to 10 mg/L).

White (1999) summarized mathematical expressions developed

by Venosa et al. (1979), the following of which are those for

unfiltered secondary effluent:

Total coliforms:

log10 TC 0.96 log10 TCI 0.89 0.012 TCOD 0.60 NO2-N

0.013 TSS 4.024 log10 T 0.57R1/2 (1)

Figure 1Reduction of fecal coliforms with transferred ozone dose for different treated effluents (and contact times):

Evry, Francesecondary; Washington, U.K.secondary; Indianapolis, Indianatertiary (Janex et al., 1999).

Table 3Transferred ozone doses for secondary

effluents corresponding to a 2-log reduction in fecal

coliforms (Janex et al., 2000).

Effluent location

Evry,

France

Washington,

U.K.

Buenos Aires,

Argentina

COD (mg O2/L) 40 15 80 15 140 30

TSS (mg/L) 47 1035 4060

Transferred O3 dose

(mg/L) 68 812 817




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viruses than chlorine (CES, 1988; Tyrrell et al., 1995). The per-

formance of ozone with other microoganisms and parasites (e.g.,

protozoan cysts and nematodes) in wastewater effluents is pres-

ently unclear because of the lack of sufficient studies. Liberti et al.

(2000) reported that in tests with tertiary-treated municipal efflu-

ents, ozone was very effective toward Pseudomonas aeruginosa,

moderately effective toward Giardia lamblia, and substantially

ineffective toward Cryptosporidium parvum, although it was

acknowledged that the low numbers of C. parvum in the un-

treated effluent made the results uncertain (Table 5). Generally,

ozone treatment of secondary effluents has been found to be

capable of achieving an effluent quality equivalent to the EU

bathing waters and WHO agricultural standards as well as other

comparable standards. However, for more stringent standards

such as the California Title 22 regulations, a tertiary-treatment

step (e.g., filtration) may be necessary before ozonation (Janex

et al., 1999).

The Effect of Ozone on Effluent Constituents

Because of its high oxidation potential, ozone reacts with a wide

range of organic and inorganic compounds in water. As previously

described, chemical oxidation by ozone occurs via two distinctreaction mechanisms, namely a molecular ozone reaction pathway

and a hydroxyl radical (OH) reaction mechanism. The molecular

reaction mechanism is more selective than the radical reaction

pathway, but the latter mechanism results in much greater reaction

rates. The predominant mechanism, or relative contribution of

each mechanism, in an ozonation process depends on various

water quality parameters such as pH, alkalinity, and organic con-

tent because these determine the presence and influence of species

that act as radical initiators, promoters, and scavengers. Many

previous ozonation studies established that ozone attacks aromatic

and unsaturated compounds, thereby affecting the chemical com-

position and the overall quality of the water (Hoigne, 1998).

Following the ozonation of four municipal secondary effluents,

Gardiner and Montgomery (1968) observed a reduction in COD,

an initial increase followed by a decrease in 5-day biochemical

oxygen demand (BOD5), and a reduction of TOC and of suspended

solids concentration. Ozone also removed 50 to 90% of anionicand nonanionic detergents. Four different types of pesticides and

five types of phenols spiked into the wastewater effluent samples

were also substantially removed after ozonation. In a pilot-plant

study using ozone for the disinfection of secondary municipal

effluent, Nebel et al. (1973) observed the removal of turbidity and

color after ozonation, and a 30% average reduction of COD. The

TOC remained primarily unaffected whereas the BOD5 decreased,

although there were cases where increases in BOD5 were ob-

served. The work of Legube et al. (1987) indicated a 25 and 50%

reduction of COD and BOD5, respectively. Langlais et al. (1992)

also found a complete or partial elimination of aromatic com-

pounds, a reduction of unsaturated fatty acids, and a 50% reduction

of anionic detergents. In the same study, the concentration ofcombined amino acids and polysaccharides decreased after ozo-

nation. Sasai et al. (1997) reported a 12% reduction of TOC and a

33% reduction of COD after ozonation of a secondary effluent.

Paraskeva et al. (1997, 1998) observed a 30% reduction in COD,

a decrease in BOD in the range of 10 to 60%, a reduction in color

between 25 and 55%, and a 254-nm reduction in UV absorbance

between 15 and 40%, whereas parameters such as TOC and pH

remained unaffected.

In terms of color reduction, ozone is known to be particularly

Table 6Summary of COD reduction by ozone treatment of secondary effluents.

Source

Ozone dose

(mg/L)

Initial COD

(mg/L) Final COD (mg/L)

Reduction

(%)

Gardiner and Montgomery, 1968 541 2656 Reduction of approximately 50% ozone dose n/aa

Nebel et al., 1973 1216 2040 n/a 23

Fressonnet-Chambarlhac et al., 1983 69 2035 1530 n/a

Legube et al., 1987 612 3375 n/a 25

Absi et al., 1993 14 40170 n/a 5

Toffani and Richard, 1995 20 120200 n/a 3050

Sasai et al., 1997 15 20 13 33

Paraskeva et al., 1998 2.55 2249 350

510 2040

1030 3075

a n/a not available.

Table 5Treatment of selected pathogens by ozone (15-mg O3/L dose, 10 minutes) in tertiary municipal effluents

(Liberti et al., 2000).

Pathogen

Feed CLa Feed Fa

Influent Treated Influent Treated

Pseudomonas aeruginosa (CFU/100 mL) 1800 28 800 8

Giardia lamblia cysts (count/L) 213 92 33 10Cryptosporidium parvum oocysts (count/L) 10 8 2 2

a CL clarified and F clarified and filtered.




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effective. Studies from two large municipal wastewater treatment

plants in Japan that receive strong influents from dyeing factories

(Hiromi and Tsujimoto, 1997; Sasai et al., 1997) showed that

ozonation achieved almost full decolorization of the effluents. In

the United Kingdom, the only ozonation plant treating secondary

effluent is operating principally for color reduction because of the

presence of dye waste in the raw wastewater. The plant is operat-

ing successfully with a color reduction of 30 to 80% in absorbancewavelengths from 400 to 600 nm (Churchley, 1998). A summary

of values for COD, BOD, and color removal from several studies

are presented in Tables 6, 7, and 8, respectively.

A small amount of information concerning the effect of ozone

on other effluent quality parameters has been reported in the

literature. Graham and Paraskeva (2001) observed that for a well-

nitrified secondary effluent, ammonia-nitrogen was further reduced

by 20% and the nitrate-nitrogen concentration correspondingly

increased slightly; however, the absolute magnitude of the changes

were small (1 to 2 mg/L). Absi et al. (1993) reported that ozone

had no effect on the ammonia-nitrogen concentration for a chem-

ically assisted primary treatment effluent where the ammonia-

nitrogen concentrations ranged from 9 to 14 mg/L. Instead, they

suggested that the presence of organic materials may have pro-

tected the ammonia from oxidation. In the same study, iron con-

centrations (residual coagulant) were slightly reduced, and mea-

surable concentrations of hydrogen sulfide in the effluent (0.8mg/L) were reduced to zero by ozone.

In general, the physicochemical quality of municipal wastewa-

ters increases systematically with ozone dose. Thus, both

Paraskeva et al. (1998) and Sasai et al. (1997) have reportedincreased COD reductions at higher ozone doses (Figure 2). The

average reduction for an ozone dose less than 5 mg/L was small.

For doses in the typical disinfection range (i.e., 4 to 10 mg O3/L),

the average COD reduction was 20 to 30% while, for higher ozone

doses (e.g., 20 mg/L), the average reduction was 30 to 50%.

Biochemical oxygen demand removal has also been found to

increase with ozone dose, although, as mentioned earlier, an in-

crease in BOD has been noted after a low ozone dose (Gardiner

and Montgomery, 1968). Reductions in effluent color and UV

absorbance (at 254 nm) are strongly correlated with ozone dose

(Paraskeva et al., 1998) (Table 8).

Because ozone is generated and applied in a feed-gas flow of

either air or oxygen (ozone concentration between 1.6 and 16% of

Table 7Summary of BOD5 changes after ozone treatment of secondary effluents.

Source Ozone dose (mg/L) Initial BOD5 (mg/L) Final BOD5 (mg/L) Reduction (%)

Gardiner and Montgomery, 1968 621 4 56 n/aa

80 12 5 n/a

Nebel et al., 1973 1216 1025 n/a 15

Legube et al., 1987 612 523 n/a 50

Absi et al., 1993 18 2648 n/a 4Paraskeva et al., 1998 510 530 1667


Table 8Summary of color reduction by ozone treatment of secondary effluents.

Source

Ozone dose

(mg/L) Color units Initial color Final color

Reduction

(%)

Nebel et al., 1973 1216 Alpha units 20100 n/aa 79

Fressonnet-Chambarlhac et al., 1983 69 mg/L platinum

cobalt standard

5 5 n/a

Toffani and Richard, 1995 20 UV absorbance

(abs) at 420 nm

(5 cm)

0.30.5 n/a 6080

Sasai et al., 1997 15 Japan IndustrialStandards Z-

7880 color

difference

1.4 0.2

Churchley and Upton, 1997 9.5 UV abs/cm at:

400 nm 0.095 0.040

450 nm 0.075 0.020

500 nm 0.070 0.020

550 nm 0.055 0.012

600 nm 0.040 0.010

650 nm 0.020 0.005

Paraskeva et al., 1998 2.55 UV abs at 400 nm

(4 cm)

0.150.20 1530

510 2555

1030 6070





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feed gas, by weight), and during reaction it decomposes to oxygen,

the concentration of dissolved oxygen (DO) in the treated water

therefore increases, almost to saturation levels (Paraskeva et al.,

1997). Although the level of dissolved oxygen in rivers is regu-

lated in many countries, the ozonation of effluents has the added

benefit of increasing the DO levels of the receiving waters. In one

reported case, increased receiving water DO was one of the prin-

cipal reasons ozone was chosen as the method of disinfection. This

involved two disinfection plants in Indianapolis, Indiana, where

the DO level of the White River dropped to zero after the waste-

water treatment plants were initially installed but increased to 8 to

10 mg O2/L after ozonation became operational (Rice, 1997).

Ozonation Byproducts

The reactions of molecular ozone and free radicals to a wide

range of organic and inorganic compounds present in natural and

contaminated waters have been well-researched in the last 20 years

and these have been reviewed elsewhere (see, for example,

Hoigne, 1998). In general, ozone is capable of reacting with a wide

range of such compounds, leading to the formation of reaction

intermediates and stable byproducts. Thus, the ozonation of sec-

ondary wastewater effluents leads to a change in the nature of the

organic matter as observed by a shift in the molecular size distri-

bution from large to smaller fractions (Paraskeva, 1998; Sasai et

al., 1997). Previous studies of the ozonation of surface waters

indicated that the most common ozonation byproducts are short-

chain aldehydes comprising up to nine carbon atoms (Glaze et al.,

1989; Schechter and Singer, 1995). Based on samples of a sec-

ondary effluent after ozone disinfection, Legube et al. (1987)

reported increases in the presence of short-chain fatty acids, car-

boxylic acids, alcohols, alkanes, ketones, and free amino acids.

Langlais et al. (1992) also reported an increase in the concentration

of aldehydes, especially heptanal and nonanal, and an increase in

short-chain saturated fatty acids (6 to 9 carbon atoms) after ozo-

nation of a secondary municipal effluent. Most recently, Liberti et

al. (2000) reported that in tests with two different tertiary-treated

municipal effluents (i.e., coagulationsedimentation and coagula-tionsedimentationfiltration) that had similar organic contents (7

mg/L dissolved organic carbon), a numerically similar increase in

total aldehydes was found (from 0.1 mg/L to 0.35 mg/L, as

fomaldehyde), with an ozone dose of 15 mg/L.

A significant concern associated with ozone in drinking water

treatment is the potential formation of halogenated substances such

as bromate, a possible carcinogen, and brominated organics arising

from the reaction of ozone and bromide. The proposed reaction

involves the intermediate formation of hypobromous acid, which

then reacts with more ozone to create bromate or reacts directly

with organic precursors to form bromoform and other brominated

species. Numerous studies have been published in the last 10 years

(e.g., Von Gunten and Hoigne, 1994) regarding factors affectingthe formation of these compounds and how to avoid or minimize

bromate formation. In contrast, the potential formation of bromi-

nated compounds in the field of wastewater treatment has received

comparatively little research attention. Among the few studies

conducted, the study by Liberti et al. (1999) did not observe any

appreciable bromate or bromoform formation after ozonation of a

secondary effluent. Although the effluent contained a significant con-

centration of bromide (3 mg/L), the lack of any measurable forma-

tion of bromate and bromoform was explained by the presence of

ammonia (22 to 23 mg/L), which reacts rapidly with hypobromous

acid to produce monobromamine; the reaction rate of hypobromous

acid with ammonia is several orders of magnitude greater than the

bromate or bromoform reactions (Von Gunten and Hoigne, 1994).

Figure 2Variation of COD with transferred ozone dose for a secondary ef fluent (Paraskeva et al., 1998).




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Strategic Diagnostics, Inc., Newark, Delaware) bioluminescence

test methods. Paraskeva et al. (1998) used the Microtox test and

reported that both unozonated and ozonated effluents were not

toxic and that a value for the 50% effective concentration factor

could not be established. In contrast, the study by Nebel et al.

(1973) found that a variety offish exposed to ozonated secondaryeffluent survived the duration of the pilot study, whereas many fish

did not survive exposure to the undisinfected effluent. Similarly,

Collivignarelli et al. (2000) showed that ozone treatment increased

the survival of test fish (Salmo gairdnerii) in diluted secondary

effluent (100% survival compared with 90% for untreated efflu-

ent). A much earlier study by Venosa and Ward (1978) reported

that ozonation reduced the toxicity of an initially toxic secondary

effluent. Biological monitoring downstream of the discharge of the

Leek Wastewater Treatment Plant (Straffordshire, England) to the

River Churnet in the United Kingdom showed that ozonation did

not induce toxicity but improved the biological health of the river,

as indicated by an increase in the biotic score (Churchley and

Upton, 1997).

Changes in effluent mutagenicity were found to be site-specific

according to Jolley et al. (1982). Results from the ozonation of

nine different effluents showed that mutagenicity could both de-

crease and increase. Greene and Stenstrom (1994) concluded that

ozonation did not change the mutagenicity of physicochemically

treated municipal effluent. Although a 7 to 24% reduction of

mutagenicity was recorded, this result was not considered statis-

tically significant. In contrast, studies by Nakamuro et al. (1989)

and Ono et al. (1992) showed significant reductions of mutagenic-

ity following the ozonation of municipal river water and of an

activated-sludge municipal effluent. Similar results are reported by

Ono et al. (1995), after the ozonation of night soil effluent, and

Sasai et al. (1997), for a wastewater secondary effluent (14 mg

O3/L). Paraskeva et al. (1999) also reported that ozone did not

induce any mutagenicity in a secondary municipal effluent, and

there was an indication that ozone could reduce the mutagenicity

of the effluent. In contrast, Collivignarelli et al. (2000) observed

that, using the Ames method, ozone at low doses (2.5 to 3 mg

O3/L) produced a low level of mutagenicity in samples of second-

ary effluent taken in both summer and winter; no mutagenicity was

recorded in untreated effluent samples. However, the results of a

second genotoxicity test, the onion (Allium cepa) roots test, did not

show any adverse effect.

Ozone-Contacting Conditions

The ozonation of wastewater effluents is typically via an

ozone gasliquid contacting system in which a nominal ozone

dose is applied in a given contact time. The highly variable

nature of wastewater flows has led to design flowrates for

ozone-disinfection systems that are typically 2 to 3 times the

average daily flowrate; additionally, the ozone-generation ca-

pacity is that required to achieve a specified applied ozone dose

at the peak design flowrate (U.S. EPA, 1986). In traditional

disinfection practice (e.g., in chlorination and the use of ozone

in drinking water disinfection), the concept of the C tproductis normally applied, where C is the concentration of dissolved

(residual) ozone measured at the outlet of the contact chamber

(in mg/L), and t is the contact time between the residual of

ozone and water (in minutes). This is a conservative approach

that is based on the well-known ChickWatson model

ln Nt/N0 kCnt (3)

where Nt

and N0 are the number of microorganisms surviving at

time t (t 0); C is the residual disinfectant; and k and n areempirically derived constants.

Normally, the value of n is assumed to be unity, and values for

the C t product corresponding to different degrees of microor-

ganism inactivation have been established from well-controlledlaboratory and pilot-plant tests (e.g., Table 2).

In applying eq 3 to wastewater disinfection there are difficulties

in that the C t approach assumes that an aqueous ozone con-centration is available throughout the contact time, t. However, a

characteristic of the high-redox potential of ozone (E0H 2.07 V)

is that, when it is applied to water, there is an instantaneous ozone

demand (occurring in the initial 20 seconds) (Roustan et al., 1997)

arising from very rapid reactions with a range of inorganic and

organic species. The magnitude of this ozone demand varies with

the nature of the water. For wastewaters, Janex et al. (1999)

reported immediate ozone demands of 3.1 to 4.2 mg/L and 7.4 to

9.6 mg/L for two secondary effluents and 2.5 to 5.3 mg/L for a

tertiary effluent. Associated with this immediate ozone demand isa significant reduction in microbiological counts (e.g., 2-log re-

duction in fecal coliforms) (Janex et al., 2000). However, there

seems to be no correlation between the extent of inactivation and

the immediate ozone demand (Janex et al., 1999). Once the im-

mediate ozone demand is satisfied, further applied ozone will

produce a measurable residual that is available for further micro-

bial disinfection. The appearance of a residual ozone concentration

(CR

) has been modeled for the case of a bubble-column reactor as

a simple equation in terms of the overall transferred (or absorbed)

ozone dose (TOD), the immediate ozone demand (X), the hydrau-

lic residence time offlow in the reactor (), and a kinetic constant,k, (Roustan et al., 1997) as follows:

CR TODX / 1 k

(4)

In theory, the utility of this model is that, provided represen-

tative values for Xand kcan be obtained for a given wastewater,

the total consumption of ozone can be determined correspond-

ing to a set ozone residual at the reactor outlet ( CR

); hence, the

ozone gas flow rate can be determined for a given ozone gas

concentration.

The principal factor that determines the disinfection perfor-

mance has been found to be the overall transferred, or absorbed,

ozone dose (in milligrams per liter). This is simply defined as the

difference in the ozone applied in the feed gas and the ozone lost

in the exhaust gas per unit volume of wastewater. Previous studies

have shown an empirical relationship between the logarithm of

Table 9Bioluminescence toxicity of urban wastewater

samples (Monarca et al., 2000).

Treatment

ECF50a

Summer Winter

Treated wastewaterb 63.0 (60.865.2) 85.7 (82.888.8)

Ozone (3 mg/L summer,2.5 mg/L winter)

16.5 (14.618.6) 205.7 (192.8219.5)

a ECF50 Effective concentration factor that inhibits 50% of the

light emitted by bacteria after 15 minutes of exposure (low

values indicate high toxicity).b Wastewater treated with activated sludge.




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bacterial survival (log10

Nt/N

0) and either the TOD or the loga-

rithm of TOD. In the case of the latter, the U.S. EPA has suggested

the following relationship of the form (U.S. EPA, 1986):

log N/N0 n log TOD/q (5)

where n is the slope of the doseresponse curve and q is the

x-axis intercept of the doseresponse curve (which is the

amount of ozone transferred before a measurable kill is ob-

served). An example of this relationship is shown in Figure 4,

which has been plotted from the data of laboratory ozonation

tests with a U.K. nitrified, secondary effluent (Paraskeva and

Graham, 1999). These results gave values for n and q of3.65and 0.77, respectively, for E. coli and 3.04 and 0.61, respec-tively, for total coliforms. These values are typically of a

similar magnitude to those obtained from field-scale trials of

nitrified secondary effluents at two plants in the United States,

where n and q were in the range of2.51 to 3.14 and 0.50 to1.05, respectively, for total coliforms, and 3.15 and 0.79,respectively, for fecal coliforms (U.S. EPA, 1986). However, in

all of these studies, the data were highly variable (Figure 4) and

the correlation coefficients (R2) are typically poor.

Contact time is an important parameter in ozone treatment

because it affects the size of the contact tanks, the treatment rate,

and the associated equipment and facilities of an ozonation plant.

Although most U.S. plants operate with a contact time of 10 to 15

minutes, there are studies indicating that efficient disinfection can

be achieved regardless of the contact time. Nebel et al. (1973)

reported that disinfection occurred after only a few seconds but, for

practical reasons, suggested a 5-minute contact time. Finch and

Smith (1989) observed no changes in the doseresponse curve of

E. coli for contact times from 1 to 22 minutes, and, most recently,

Janex et al. (1999) reported that for a given TOD, a 2-minute

contact time achieved the same fecal coliform inactivation as 10

minutes. More studies are necessary to confirm whether this be-

havior generally applies to all types of microorganisms of interest

because such microorganisms have a range of tolerances to ozone.

The need to reduce the comparatively high costs of ozone

technology in the last 10 years has resulted in the production of

more efficient ozone corona discharge generators capable of pro-

ducing ozone at concentrations of up to 6% w/w in air and up to

20% w/w in oxygen. These generators are able to reduce energy

costs by producing more ozone at approximately the same energy

consumption level as traditional generators, to economize oxygen

consumption, increase the mass-transfer rate of ozone with the

appropriate contact chamber, and reduce the treatment rate (Dyer-

Smith, 1997). Such generators not only provide a faster treatmentbut also significantly reduce the size of contact chambers. In the

study by Paraskeva et al. (1998), three gaseous ozone concentra-

tions (1.5, 4.5, and 13.5% w/w in oxygen) were studied with

respect to their effect on the treatability of the secondary effluent.

It was found that neither the ozone concentration nor the contact

time were important factors as far as the overall treatment (in terms

of the principal quality parameters COD, BOD, TOC, color, etc.)

was concerned. The single factor governing the extent of treatment

was found to be the TOD, as shown in Figure 5. However, higher

gaseous inlet ozone concentrations resulted in a faster treatment

(Figure 6). In view of these results, the optimal design of the

ozone-contacting system should be based on maximizing the

ozone transfer rate, corresponding to an overall TOD that achievesthe required degree of effluent treatment.

Conclusions

The use of ozone for wastewater disinfection has received a

considerable amount of attention in the last 20 years. Interest in

this application of ozone is likely to increase in the future, partic-

ularly where wastewater effluents are to be reused and a high

degree of treatment is required. Given its high reactivity with a

wide range of substances and relative cost of application, ozone is

likely to be most appropriate for achieving specific treatment goals

with effluents that have received some degree of pretreatment,

such as secondary effluents. Thus, ozone treatment of secondary

effluents has been found to be able to meet bacteriological stan-

Figure 4Reduction of E. coli and total coliforms with transferred ozone dose (Paraskeva and Graham, 1999).




11/13

dards typically required for discharge to recreational waters, and a

high degree of virus removal has been reported. Limited data are

currently available on the inactivation of specific pathogenic or-

ganisms (e.g., protozoa cysts) in wastewater effluents or their

regrowth potential following ozonation. Given the highly variable

nature of effluent quality, the performance of ozone treatment islikely to be equally variable, and thus considerable attention will

need to be given to plant design and operation of the ozone system.

Attempts to model bacterial removal during ozone treatment have

been frustrated by the inherent complexity of the process, although

some empirical performance relationships exist for the principal

indicator bacteria. The TOD has been found to be a signi ficant

parameter that determines microbiological performance. Clearly, a

part of the TOD is represented by the immediate ozone demand

and a significant degree of microbial inactivation occurs by satis-fying this immediate demand. Further studies would be useful to

develop a process model that can represent the simultaneous

consumption of ozone by microbial inactivation and the effluent

Figure 5Variation of effluent of COD with transferred ozone dose at different gaseous ozone concentrations (4.5 and

13.5 % in oxygen, w/w; constant ozonation rate) (Paraskeva et al., 1998).

Figure 6Effect of gaseous inlet ozone concentration (20 and 60 mg/L) on E. coli inactivation (Paraskeva and Graham,

1999).




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Ozonation of Municipal Wastewater Effluents