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Ruqiao Shen

A thesis submitted in conformity with the requirements
for the degree of Doctor of Philosophy

Graduate Department of Civil Engineering
University of Toronto

© Copyright by Ruqiao Shen (2013)

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Ruqiao Shen

Doctor of Philosophy, 2013
Graduate Department of Civil Engineering
University of Toronto


N-nitrosamines are considered as a group of emerging disinfection byproducts (DBPs) with

potential carcinogenicity at ng/L level. The presence of nitrosamines in drinking water is most

commonly associated with chloramination of amine-based precursors. This research investigates

the potential of amine-based pharmaceuticals and personal care products (PPCPs) as nitrosamine

precursors under practical drinking water disinfection conditions, as well as some critical factors

that may affect the nitrosamine formation via PPCPs.

All of the twenty selected PPCPs were able to form the corresponding nitrosamines upon

chloramine disinfection, and eight of them rendered molar conversions higher than 1 % under

practical disinfection conditions. Ranitidine had the highest N-nitrosodimethylamine (NDMA)

molar conversion among the tested PPCPs.

A three-parameter kinetic model was proposed to describe and predict the NDMA formation

from pharmaceuticals during chloramination in various water matrices. The model accurately

reflected all three significant characteristics of the NDMA formation curve, including an initial

lag phase, followed by a fast increase in NDMA formation, and eventually reaching a plateau.

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2006); Otonabee River, < 11.4 µg/L (Woodbeck, 2007)), while bromide concentrations applied

were much higher (mg/L) for studies that have reported these effects, so the differences in the

observed NDMA formation profiles were thought to be due to some aspect of the NOM. Since

bromide is in higher concentration and so may be more of a concern in coastal waters due to

saltwater intrusion, the potential impact from bromide was considered to be outside the scope of

the present tests but would be of interest for future study on NDMA formation in coastal waters.

NOM may affect NDMA formation in two ways: competition for chloramine and direct

interaction with PPCPs. The influence of NOM’s competition for chloramine was considered to

be minimal due to the small observed chloramine decay (less than 50 % chloramine decay during

the course of the kinetic experiments, see Appendix 2, Figure A2.2) and the large excess of

chloramine relative to the pharmaceuticals (mg/L vs. lower µg/L) at the end of the kinetic

experiment. On the other hand, NOM may interact with the pharmaceuticals and then inhibit

their reaction to form NDMA, and it is these interactions that were thought to better explain the

observed results. Previous studies have demonstrated that aromatic amines undergo reversible

covalent binding with carbonyls and quinones in soil humic substances in the environment

(Parris, 1980; Thorn et al., 1996; Weber et al., 1996). For example, Weber et al. (1996) have

observed significant binding at a concentration of 466 µg/L of aniline and 250 mg/L of humic

substances, i.e., 1.86 µg/mg sorption ratio; while in this work, 6 ~ 8 µg/L of selected

pharmaceuticals were dosed into 2 ~ 6 mg/L of TOC (the majority of which was humic

substances, see details in Section 3.1.3). Therefore, the potential sorption ratio was relatively

comparable with the study by Weber et al (1996). Considering the trace level of pharmaceuticals

relative to the amount of NOM in natural water samples, a similar scenario with the sorption of

anilines onto soil/sediment organic matter could exist. Therefore, it is possible that certain

fractions or functional groups in NOM may interact with these amine-based pharmaceuticals and

thus hinder their initial contact with chloramine species. As the binding is reversible and

chloramine is in large excess, eventually the NDMA conversion from pharmaceuticals can still

reach the maximum level given enough reaction time.

Currently, although no direct spectroscopic evidence exists for the NOM-pharmaceutical binding

in the aqueous phase, this theory is indirectly supported by some literature investigating the

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removal of pharmaceuticals during coagulation/flocculation processes, where the removal of

pharmaceuticals was likely due to the sorption or electrostatic attraction onto particulate organic

matter and co-removed through the settling process (Ballard and Mackay, 2005; Diemert, 2012;

Stackelberg et al., 2007; Vieno et al., 2006; Westerhoff et al., 2005). Stackelberg et al. (2007)

also detected the target pharmaceuticals in the dried solids of settled sludge. In addition, de

Ridder et al. (2011) observed enhanced removal of some positively charged pharmaceuticals

using granular activated carbon preloaded with NOM. They attributed the enhancement to the

electrostatic attraction since the surface of NOM is usually negatively charged due to the

abundance of carboxyl groups. In the current study, the selected amine pharmaceuticals are

positively charged at neutral pH, therefore this possible electrostatic attraction may also lead to

the formation of NOM-pharmaceutical complexes.

The proposed NOM-pharmaceutical binding theory can well explain the initial lag phase

observed in the kinetic study, thus theoretically the length of the initial lag phase should be

proportional to the amount and/or type of NOM in the water samples. NOM components can be

at least partially described by the samples’ TOC and SUVA values. It was observed that water

with higher TOC and SUVA levels tended to have a longer initial lag phase; all four

pharmaceuticals exhibited their longest initial lag period in river water samples. However, the

treated river water had higher TOC and SUVA values than the lake water, yet ranitidine showed

a slightly shorter initial lag phase in the treated river water (Figure 5.2). According to the

literature, binding usually occurs between aromatic amines and some specific functional groups

in the humic substances; therefore, it is the amount of these functional groups that directly affects

the binding, rather than the amount of NOM. The treated river water and lake water came from

different sources, thus they could have different compositions of surface functional groups. In

addition, the treatment processes at the water treatment plant, especially the

coagulation/flocculation process, not only remove NOM, but can also change its surface

characteristics (e.g., steric structure, charge distribution, etc.). This suggests that some specific

NOM fractions or moieties might be more relevant than would be indicated by simple bulk

measurements of water quality, such as TOC and SUVA.

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In addition, in Chapter 7, Section 7.3.2, it is suggested that the higher TOC level in the river

water might be responsible for the “rebinding” between ranitidine and the NOM breakdown

products which was not observed in the lake water with a lower TOC. The LC-OCD results

indicate that upon the sequential disinfection, humics was first degraded into building blocks,

and further breakdown of building blocks would lead to the increase in LMW fractions. There is

no significant difference in the distribution of NOM fractions between the two matrices,

although the river water has a higher portion of humics while the lake water has a higher portion

of biopolymers (Figure A15.3). However, considering the much higher TOC level in the river

water, the absolute amount (in mg/L organic carbon) that changed upon the treatment process

can be much bigger than the lake water.

Figure A15.3. The distribution of hydrophilic NOM fractions for two water matrices

Figure A15.4 compares the absolute amount of each NOM fraction changed (in mg/L organic

carbon) along the sequential chlorination (120 min) and chloramination treatment in the two

matrices. In general, biopolymers and humics were both degraded, but the reduced amount was

much bigger in the river water. In the lake water, the reduction in building blocks led to the

increase in LMW fractions; while in river water, the building blocks increased first (due to the

breakdown of humics) and then degraded into LMW fractions. The increase in building blocks

and LMW fractions were also bigger in the river water than that in the lake water. As such, there

are more NOM breakdown products in the river water due to the breakdown of large NOM

fractions, which may be responsible for the “rebinding” with ranitidine and thus the further

inhibited NDMA conversion as chlorine contact time increased.

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Figure A15.4. Comparison of the change of NOM fractions upon sequential chlorination (120 min) and
chloramination in Lake Ontario water and Otonabee River water

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