Introduction

Organic pollutants in municipal wastewater consist of a wide range of synthetic chemicals, such as antioxidants, plasticizers, pharmaceuticals and personal care products. These pollutants are biologically active to some extent and pose a threat to the aquatic environment and human health (Daughton and Ternes 1999). To minimize the potential risk, efforts are underway to reduce the exposure of pollutants in waters. However, conventional wastewater treatment technologies, such as activated sludge treatment and sorption, do not necessarily achieve high removal efficiency for these emerging pollutants, especially trace organic chemicals, since they are hydrophilic, less volatile, and exhibit toxic effects to microorganisms (Oulton et al. 2010; Xiao et al. 2014a, b, c). Therefore, effluent from municipal sewage treatment plant is the major source for trace pollutants to enter natural waters.

Ultrasound is an emerging advanced oxidation process that can quickly and effectively degrade organic pollutants in waters. It has several unique advantages when compared to other conventional oxidation technologies, including lack of potentially harmful chemicals, ease of use, and short contact time (Adewuyi 2005a, b; Hoffmann et al. 1996). When water is exposed to ultrasound, acoustic pressure waves are generated. The acoustic pressure waves consist of compression and rarefaction cycles. In the rarefaction cycle of the acoustic pressure wave, it leads to the formation of bubbles from the gas nuclei that exist in water. In the compression cycle, the bubble volume decreases due to increasing pressure in the surrounding water (Mason and Lorimer 1988). Thus, bubbles grow and shrink in response to alternating acoustic pressures. Within several compression and rarefaction cycles, bubbles collapse when the ultrasonic intensity is beyond the cavitation threshold, known as cavitation bubbles (Suslick 1989).

The collapse of cavitation bubbles causes extreme conditions that are depicted by the hot spot theory, as shown in Fig. 1a (Suslick et al. 1986). In the center of collapsing bubble (i.e., gas region), the temperature and pressure are approximately 5000 K and 1000 atm, respectively (Flint and Suslick 1991). The high temperature leads to the breakdown of gaseous water molecule in the bubble to hydroxyl radicals (·OH) and hydrogen atoms (·H) (Flint and Suslick 1991):

$$ {\text{H}}_{ 2} {\text{O}}\mathop{\longrightarrow}\limits^{(((}{}^{ \cdot }{\text{OH}} + {}^{ \cdot }{\text{H}} $$
(1)

The temperature in the interfacial region surrounding the hot core is estimated to be 1900 K, and the ·OH concentration is estimated to be up to 4 mM (Gutierrez et al. 1991). The thickness of the interfacial region is estimated to be 200 nm (Mason et al. 1990). The temperature of bulk phase surrounding the cavitation bubble is ambient. The ·OH formed in the bubble core diffuses to the bubble–water interface and then to the bulk phase due to ·OH concentration gradient, as shown in Fig. 1b (Adewuyi 2005a). Therefore, in a sonicated solution, an organic molecule undergoes degradation by two different pathways: decomposition by heat in the gas and interfacial regions of the cavitation bubbles and ·OH oxidation in the gas, interfacial, and bulk regions, as indicated in the follow equation (Adewuyi 2001; Hoffmann et al. 1996; Mendez-Arriaga et al. 2008; Riesz et al. 1985; Weavers et al. 2005):

$$ \left( {\frac{{{\text{d}}C}}{{{\text{d}}t}}} \right)_{\text{obs}} \,=\, k_{\text{thermolysis}} [C] \, + k_{{^{ \cdot } {\text{OH}}}} \left[ C \right][^{ \cdot } {\text{OH}}] $$
(2)

where \( \left( {\frac{{{\text{d}}C}}{{{\text{d}}t}}} \right)_{\text{obs}} \) is the observed degradation rate of a given organic pollutant in the unit of M min−1, k thermolysis and \( k_{{^{ \cdot } {\text{OH}}}} \) are the thermolysis and ·OH oxidation rate constants for the organic pollutant, respectively; [C] and [·OH] represent the concentrations of the pollutant and ·OH in solution, respectively.

Fig. 1
figure 1

Schematic diagrams of hot spot theory for a cavitation bubble (a) and corresponding spatial distributions of temperature (T), ·OH, and organic pollutants surrounding the bubble (b; vertical axis is arbitrary). In the hot spot theory, the cavitation bubble is divided into three regions: gas phase, interfacial zone, and bulk phase. The temperature decreases rapidly from as high as 5000 K at the bubble core to ambient in the bulk solution. ·OH is generated from thermolysis of water molecules in the gas phase and diffuses to the interface and bulk solution due to the concentration gradient. Organic pollutants undergo two major degradation pathways, thermolysis and radical oxidation, thereby exhibiting a increasing trend in spatial distribution as compared to temperature and ·OH

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In the past three decades, intensive efforts have been exerted to understand various factors determining the effectiveness of ultrasound in removing organic pollutants from waters. In general, three factors affect the removal efficiency of organic pollutants in waters, ultrasonic operational conditions (e.g., ultrasonic intensity, frequency, and mode), physicochemical properties of organic pollutants (e.g., surface excess, Γ, octanol–water partition coefficient, K OW, Henry’s law constant, K H, and diffusivity, D), and solution chemistry (e.g., ionic strength and pH) (Adewuyi 2005a, b; Bolong et al. 2009; Brotchie et al. 2009; Colussi et al. 1999; Francony and Petrier 1996; Jiang et al. 2002; Mendez-Arriaga et al. 2008; Mizukoshi et al. 1999; Nanzai et al. 2008; Pee et al. 2015; Petrier et al. 1996; Xiao et al. 2013b). Studies on the impacts of solution chemistry on sonochemical degradation of contaminants have focused mainly on the pH (Chakinala et al. 2007; Ince et al. 2009; Jiang et al. 2002; Tauber et al. 2000; Uddin and Hayashi 2009), since it is considered the most practical and easiest adjustable parameter during treatment processes. These studies are vital for a better understanding of the application of ultrasound in drinking water and wastewater treatments, since water sources usually exhibit different pH values, depending on the geological conditions of groundwater or the origins of wastewaters (i.e., industrial or domestic wastewaters).

Although pH exerts a significant influence on removal efficiency of organic pollutants, the influence is not always consistent and there is lack of systematic investigation on the pH effect. The rationale for this review is to provide a comprehensive review to fulfill the knowledge gap in pH effects on sonochemical degradation of organic pollutants. In particular, we conducted a systematic literature investigation on pH effects, covering a temporal range from 1990s to 2010s, totaling 78 peer-reviewed literature, see Table 1. Among the list, combined systems such as ultrasound-UV, ultrasound-ozone, and ultrasound-Fenton are also included, as well as studies that include no or unclear explanations on the pH effects (Cyr et al. 1999; Goel et al. 2013; Goskonda et al. 2002; Hua et al. 1995; Okouchi et al. 1992; Shimizu et al. 2007; Shriwas and Gogate 2011; Suri et al. 1999; Wang et al. 2008; Wu et al. 1992, 2001b; Yoo et al. 1997; Zhang et al. 2007; Zhou and Ma 2006). We discuss how pH alters cavitation effects including collapse temperature and ·OH formation, followed by pH-induced changes in physicochemical properties for organic pollutants, effects of coexisting species in water (i.e., water matrices), and consequent degradation kinetics of organic pollutants in the sonication system. To the best of our knowledge, this is the first summary of pH effects on the sonochemical degradation kinetics of organic pollutants.

Table 1 Literature summary of pH effects on sonochemical degradation of organic pollutants
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pH effects on collapse temperature, radical yield and reactivity, and consequently degradation kinetics of organic pollutants

Collapse temperature

According to Eq. 2, collapse temperature of a cavitation bubble (i.e., k thermolysis), production of free radicals (i.e., [·OH]), and their distribution surrounding the bubble all impact organic pollutant degradation kinetics to different extents (Fig. 1b). At collapse, the maximum temperature (\( T_{ \hbox{max} } \), K) for cavitation bubbles is expressed as follows:

$$ T_{ \hbox{max} } = \, T_{0} \left[ {\left( {K - 1} \right)\frac{{P_{m} }}{P}} \right] $$
(3)

where T 0 is the ambient temperature (K), K is the specific heat ratio (unitless), P m is the pressure in liquid upon bubble collapse (Pa), and P is the pressure inside of the bubble at its maximum size (Pa). According to Eq. 3, the collapse temperature of cavitation bubbles depends on the saturation of different gases into solution (i.e., K) and ultrasonic operational parameters such as intensity and frequency that will alter P and P m . The pH indirectly affects the collapse temperature of cavitation bubbles (case #1 and 2 in Table 1). For instance, Drijvers et al. (1996) observed that CO2 formed from mineralization of trichloroethylene during sonication diffuses into the gas phase at acidic pH, thereby reducing the specific heat ratio of vapor and thus the collapse temperature. They also attributed the unchanged sonochemical degradation kinetics of chlorobenzene at different pH to the low CO2 yield, thus fixing the specific heat index and consequent collapse temperature of cavitation bubbles (Drijvers et al. 1998).

Radical yield

For the radical oxidation, the degradation kinetics of organic pollutants are controlled by both radical reactivity (i.e., \( k_{{^{ \cdot } {\text{OH}}}} \)) and quantity (i.e., [·OH]), as shown in Eq. 2. Many studies (case #3–18 in Table 1) observed decreasing degradation kinetics of organic pollutants with increasing pH and cited three explanations for the ·OH yield: (1) ·OH self-combines to form H2O2 at high pH; (2) ·OH is scavenged by the buffer solutions, such as \( {\text{HCO}}_{3}^{ - } \) and \( {\text{CO}}_{3}^{ 2- } \) at high pH; and (3) ·OH deprotonates at pH > 11 (Adewuyi and Appaw 2002; Ghodbane and Hamdaoui 2009; Ince and Tezcanli-Guyer 2004; Mendez-Arriaga et al. 2008; Svitelska et al. 2004; Uddin and Hayashi 2009; Yim et al. 2002). However, in synergistic systems combining ultrasound, UV, ozone, and H2O2, the effect of pH on radical yield and consequent organic pollutant degradation kinetics is more complex. Particularly, ultrasound-ozone combination exhibited increased degradation kinetics with an increase in pH (He et al. 2007a, b; Quan and Chen 2011; Sierka 1984). Elovitz et al. (2000) attributed the enhanced degradation kinetics to the OH− initiated decomposition of ozone and possible high ·OH yield at high pH in the ultrasound-ozone system. The ultrasound-UV–H2O2 system has the same trends as seen in the ultrasound-ozone system (Fung et al. 2000; Poon et al. 1999). In the combined system of ultrasound-Fenton, acidic condition was favorable for degradation since dissolved iron facilitates ·OH formation at low pH through Fenton reaction, while iron concentration is quite low at high pH (Cai et al. 2016a, b; Katsumata et al. 2011; Liang et al. 2007).

On the other hand, different free radical formation and reaction pathways at various pH result in different quantities of oxidative radicals during sonication, thereby altering the degradation kinetics of organic pollutants (Adewuyi and Appaw 2002; Bielski et al. 1985; Buxton et al. 1988; Czapski and Dorfman 1964; Ross and Ross 1977; Fung et al. 2000; He et al. 2007a, b; Kotonarou et al. 1992; Matheson and Rabani 1965; Orzechowska et al. 1995; Pang et al. 2011; Poon et al. 1999; Sehested et al. 1968; Sierka and Amy 1985; Svitelska et al. 2004; Weavers et al. 2000; Yu 2004). However, very few studies have quantitatively investigated oxidative radicals at different pH during sonolysis. Therefore, it is necessary to revisit this issue and scrutinize the potential influence of free radicals (e.g., hydroperoxyl radical, ·HO2 −) on organic pollutant degradation.

Radical reactivity

The reactivity of ·OH (i.e., \( k_{{^{ \cdot } {\text{OH}}}} \) in Eq. 2) is also used to explain the sonochemical degradation trend of organic pollutants at different pH values (case #19–27 in Table 1). For example, different degradation rates of non-ionizable chemicals were attributed to the higher oxidation potential of ·OH in acidic solutions (E° = 2.78 V) than that in neutral and basic solutions (E° ≤ 1.80 V) (Park et al. 2000). The oxidation potential of ·OH was also considered to contribute the varied degradation kinetics of rhodamine B and methyl parathion in a pH-dependent manner in ultrasonic systems (Patil and Gogate 2012; Wang et al. 2009). Nakui et al. (2007) correlated hydrazine degradation kinetics to reaction rate constants at different pH values, confirming varied ·OH reactivity in the presence H+ and OH−.

pH effects on physicochemical properties of target organic pollutants and their degradation kinetics

In Eq. 2, the proportion of each pathway (i.e., k thermolysis or \( k_{{^{ \cdot } {\text{OH}}}} \)) to the whole degradation is dependent on the physicochemical properties of organic pollutants (Adewuyi 2001; Hoffmann et al. 1996; Xiao et al. 2014b). Usually, pH affects degradation kinetics of non-ionizable organic pollutants by altering quantity and reactivity of free radicals. For acidic or basic organic pollutants, pH controls the protonation of organic acids and bases and thus their degradation kinetics, as shown in Fig. 2. In particular, the pH-dependent protonation or deprotonation results in changes in volatility of a compound described by Henry’s law constants (\( K_{\text{H}} \)) (Ashokkumar et al. 1999, 2000; Guo et al. 2005; Ku et al. 1997; Lin et al. 1996; Lin and Ma 1999; Price et al. 2002; Singla et al. 2004; Sivakumar et al. 2002; Tauber et al. 2000), hydrophobicity described by octanol–water partition coefficient (K OW) (Behnajady et al. 2008; Chakinala et al. 2007; Chen and Huang 2011; Cost et al. 1993; Dalhatou et al. 2013; Gultekin and Ince 2008; Ince et al. 2009; Okitsu et al. 2005, 2008; Ozen et al. 2005; Peller et al. 2001; Saharan et al. 2012; Serpone et al. 1992; Shemer and Narkis 2005; Song et al. 2006; Vajnhandl and Le Marechal 2007; Wang et al. 2007; Weavers et al. 2005; Wu et al. 2001a; Yang et al. 2013), and Coulombic interactions between the compound and the cavitation bubbles (Anju et al. 2012; Cheng et al. 2010; De Bel et al. 2009; Jiang et al. 2002; Kaur and Singh 2007; Kidak and Ince 2006; Kim et al. 2001).

Fig. 2
figure 2

Speciation, location, and degradation of a hypothetical acidic organic compound (HOC = OC− + H+) from acidic to alkaline condition in a sonicated solution. Note the HOC in neutral form at acidic pH accumulates on the bubble–water interface and diffuses into bubble core resulting in intensive radical oxidation and thermolysis reaction, while OC− in deprotonated form at alkaline pH remains in the bulk solution, where radical oxidation is the major reaction pathway

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Volatility

The volatility of an organic compound in aqueous solution is described by the Henry’s law constant of K H (Pa m3 mol−1), which is a measure of the partition of a compound between the gas and water phases defined by Eq. 4:

$$ K_{\text{H}}^{{}} = \frac{p}{ [C ]} $$
(4)

where p is the partial pressure of target compound in the aqueous solution (Pa). Similar to any other thermodynamic properties, K H is dependent on temperature, as indicated by Eq. 5:

$$ K_{\text{H}} = K_{\text{H}}^{{\circ }} { \exp }\left( {\frac{{ -\Delta _{\text{soln}} H}}{\text{R}}\left( {\frac{ 1}{T} - \frac{ 1}{{T^{{\circ }} }}} \right)} \right) $$
(5)

where \( \Delta _{\text{soln}} H \) is the enthalpy of solution (J mol−1), R is the universal gas constant (J K−1 mol−1), T° is the standard temperature (K), and \( K_{\text{H}}^{ \circ } \) is the Henry’s law constant at T°. Therefore, temperature control for the bulk solution is critical for sonolysis experiments to minimize the changed volatility of an organic compound due to the temperature increase during sonication.

Many studies have investigated the influences of K H on sonochemical degradation kinetics (Ayyildiz et al. 2007; Colussi et al. 1999; De Visscher 2003; Nanzai et al. 2008; Petrier et al. 1998, 2010). The ionized form of a volatile or semi-volatile compound (e.g., phenol) at higher pH remains in the bulk phase during sonication, but at lower pH the neutral species with high Henry’s law constant diffuses onto the bubble–water interface and evaporates into the gaseous phase, where intensive thermolysis and oxidation occur (Mason and Tiehm 2001; Mason and Lorimer 2002; Suslick 1989, 1990; Suslick et al. 1986). As shown in Table 1 (case #28–37), a decrease in degradation rates was observed for volatile or semi-volatile organic pollutants, such as chlorophenol and nitrophenol, with an increase in pH at various ultrasound frequencies (20–1000 kHz) and purging gases (i.e., air, argon and oxygen). All pK a values fall into the tested pH range in each study to assure the protonation state of the organic pollutants. Since not all organic molecules diffuse into gas phase of cavitation bubbles, hydrophobicity-induced accumulation of organic pollutants onto bubble–water interface also accounted for the enhanced degradation for neutral form species as compared to the ionic species.

Hydrophobicity

Octanol water partition coefficient, K OW, is a physicochemical property that is a measure of the hydrophobicity of a compound (McNaught and Wilkinson 2000), as indicated in the following equation:

$$ K_{\text{OW}} = \frac{{ [C ]\,{\text{in}}\,{\text{octanol}}\,{\text{phase}}}}{{ [C ]\,{\text{in}}\,{\text{water}}\,{\text{phase}}}} $$
(6)

Many studies have evaluated the effect of K OW on the sonolytic degradation of organic contaminants (Emery et al. 2005; Fu et al. 2007; Nanzai et al. 2008; Park et al. 2011; Wu and Ondruschka 2006). Particularly, faster sonolytic kinetics of non-volatile acidic organic pollutants were observed at low pH, because the increased hydrophobicity of protonated species resulted in more molecules accumulating and degrading at the interface of cavitation bubbles, the site of reactivity (Mason and Tiehm 2001; Mason and Lorimer 2002; Suslick 1989, 1990; Suslick et al. 1986). As shown in Table 1, cases #38–57, the majority of these studies reported reduced degradation of organic pollutants in ionized form at high pH. For example, Jiang et al. (2002) studied the sonolysis of 4-nitrophenol (NP = NP− + H+, pK a = 7.08) and aniline (ANI+ = ANI + H+, pK a = 4.6) with pH ranging from 2 to 9. They observed that the degradation rate of 4-nitrophenol decreased with an increase in pH, but the degradation rate of aniline increased with an increase in pH. They attributed the faster degradation rates of neutral 4-nitrophenol and aniline over their ionic forms to the protonation of phenoxide group for 4-nitrophenol at acidic pH and deprotonation of ammonium group for aniline at alkaline pH, respectively. The higher hydrophobicity of neutral form than the ionic form results in a higher degree of accumulation at bubble–water interfaces for the neutral species. It is worth mentioning that hydrophobicity was also the predominant factor that has been frequently used to account for organic pollutant degradation in ultrasound-UV, ultrasound-TiO2, and hydrodynamic cavitation systems (Chen and Huang 2011; Wu et al. 2001a; Yang et al. 2013).

Since surfactants always stay transphilic in different pH solutions due to a polar head and hydrophobic tail structure, no impact on the surfactant sonolysis was observed (Weavers et al. 2005). On the other hand, compounds without pK a values such as trihalomethane have no ionized form and little degradation variance for trihalomethane was observed in tested pH range (Shemer and Narkis 2005). These unchanged degradation kinetics further confirms the hydrophobicity explanation for pH effects.

Charge

Watmough et al. (1992) used a 1 kV potential electrode to record the dye (i.e., methylene blue and sky blue dye) deposition on paper in a sonicated solution. Their results implied that the ultrasound-induced gas bubbles carry a negative electric charge with a field charge of about 7 × 105 V m−1. Therefore, the electrostatic attractive force between the positively charged organic molecule and the negatively charged bubble–water interface was reported to contribute to the altered degradation kinetics at different pH values (case #58–64 in Table 1). De Bel et al. (2009) reported that the sonochemical degradation rate constants for zwitterion ciprofloxacin (pK a1 = 3.64, pK a2 = 5.05, pK a3 = 6.95 and pK a4 = 8.95) was almost four times larger at pH 3.0 than those at pH 7.0 and 10.0. They explained that the electrostatic attractive force between the positively charged ciprofloxacin molecule (+3 charged at pH < 3.64; +2 charged between pH 3.64 and 5.05; +1 charged between pH 5.05 and 6.95; 0 charged between pH 6.95 and 8.95; and −1 charged at pH > 8.95) and the negatively charged bubble–water interface resulted in a faster degradation kinetics under acidic condition than neutral and alkaline pH (De Bel et al. 2009). In addition, Kim et al. (2001) stated that dibenzothiophene during sonolysis became increasingly charged at higher pH resulting in increased degradation rates.

Other researchers have come up with alternate explanations. For example, Cheng et al. (2008) monitored degradation of perfluorooctane sulfonate and perfluorooctanoate at different pH values and attributed the fast kinetics at low pH to interactions of protons with the bubble–water interface. They suggested that the bubble–water interface became increasingly positively charged as pH decreased below 4 and therefore attracted more contaminants with opposite charge (Cheng et al. 2010). Likewise, Jiang et al. (2002) and Kidak and Ince (2006) both claimed hydrophobicity and charge of 4-nitrophenol and phenol molecules as reasons for decreasing degradation with respect to pH values (Jiang et al. 2002; Kidak and Ince 2006). In sonocatalysis systems, pH altered the surface charge of metal oxides and thus adsorption of phenol and reactive red dye 198 onto the oxide surface leading to varied degradation (Anju et al. 2012; Kaur and Singh 2007).

Many of these studies have utilized combinations of explanations for organic pollutant sonolysis at different pH in the heterogeneous cavitational system. For example, Mendez-Arriaga et al. (2008) observed that the initial degradation rate at pH 3.0 was significantly higher than those at pH 5.0 and 11.0 during sonolysis of ibuprofen. They explained that the protonation of the carboxylic group of ibuprofen at pH values lower than its pK a, 4.9, led to a faster degradation rate because the increased hydrophobicity of protonated ibuprofen results in more ibuprofen molecules accumulating and degrading at the interface of the cavitation bubble (Mendez-Arriaga et al. 2008). Additionally, they also discussed that ·OH recombined to form H2O2 at high pH resulting in a low radical yield that further slowed down the degradation process of ibuprofen. Since the sonochemical process involves thermolysis, ·OH oxidation, and formation of byproducts in gas, interfacial, and bulk regions of cavitation bubbles, a single physicochemical property of organic pollutants may have limited capability to accurately govern the complex kinetics, especially when the compound covers a diversity of structures and a wide range of physicochemical properties.

pH effects on water matrices and subsequent organic pollutant degradation kinetics

In addition to the physicochemical properties of organic pollutants, solution chemistry including both inorganic and organic matrices in waters is a critical factor in determining the final degradation kinetics by altering [·OH] and [C] in Eq. 2. Generally, inorganic matrices can change the sonochemical degradation of contaminants through two mechanisms: competing for ·OH and altering the accumulation of organic compounds at the bubble–water interface through salting-out effects. Studies on effects of inorganic scavengers on organic pollutant degradation kinetics in different pH solutions are mostly focused on buffer ions, such as \( {\text{HCO}}_{3}^{ - } \), \( {\text{CO}}_{3}^{ 2- } \), \( {\text{HPO}}_{3}^{ 2- } \) and \( {\text{SO}}_{4}^{ 2- } \). At acidic conditions, \( {\text{SO}}_{4}^{ 2- } \) and \( {\text{HPO}}_{3}^{ 2- } \) are reported to scavenge ·OH (Uddin and Hayashi 2009; Xu et al. 2013), whereas \( {\text{HCO}}_{3}^{ - } \) and \( {\text{CO}}_{3}^{ 2- } \) consumed the free radical in bulk solution at alkaline pH (He et al. 2007a; Ince and Tezcanli-Guyer 2004; Wang et al. 2009). Cheng et al. (2010) investigated the effect of specific anions on sonolysis of perfluorooctane sulfonate and perfluorooctanoate at the frequency of 612 kHz. They observed the role that anions played on the degradation kinetics followed the Hofmeister series: \( {\text{ClO}}_{4}^{ - } \) > \( {\text{NO}}_{3}^{ - } \) > Cl− > \( {\text{SO}}_{4}^{ 2- } \). They speculated that the coordinating structure of water clusters at the bubble–water interface may be forced to transform due to the presence of these ions and alter water vapor transported into the bubble, resulting in a decreased collapse temperature.

On the other hand, the natural organic matters, such as Suwannee River fulvic and humic acid, have been selected as representative organic matrices to examine their influence on sonochemical degradation of organic pollutants (Cheng et al. 2008; Laughrey et al. 2001; Lu and Weavers 2002; Taylor et al. 1999; Xiao et al. 2013a, 2014a). Reduced sonochemical degradation of target contaminants, such as 4-chlorobiphenyl (4-CB) (Lu and Weavers 2002) and polycyclic aromatic hydrocarbons (Taylor et al. 1999) by natural organic matters have been reported. Both attribute the reduced degradation kinetics to two effects: (1) natural organic matters competes with target contaminants for ·OH, potentially hindering the degradation process; and (2) Suwannee River fulvic acid alters the threshold of transient cavitation via surface tension (γ, J m−2) changes, reducing the bubble–water interfacial temperatures and ultimately degradation rates. However, some studies suggest that the presence of natural organic matters has no impact on the sonochemical degradation of contaminants, such as methyl tert-butyl ether (Kang et al. 1999). Examining the pH effects on sonochemical removal of organic pollutants in the presence organic matrices is a more complex process, since any change in a single parameter results in an alteration in both target compounds and organic matrices. So far, there is a lack of systematic studies exploring the impact of organic matrices on organic pollutant sonolysis under different pH conditions.

Conclusion

In the present review, several common explanations were summarized for the pH effects on sonochemical degradation: volatility and hydrophobicity of the target compound, Coulombic interactions between compound and cavitation bubble, and radical quantity and reactivity in different pH solutions, as tabulated in Table 2. The physicochemical property, pK a value of ionizable organic pollutants seems to be one of the most critical factors in determining sonochemical degradation kinetics at different pH. Volatility, hydrophobicity and Coulombic interactions are the most widely used explanations for sonochemical degradation of organic pollutants with pK a values. However, these explanations are not sufficient to explain the degradation variance at different pH for organic pollutants without ionizable groups (e.g., PAHs and carbamazepine), since they do not protonate or deprotonate with pH. Instead, the oxidation potential and quantity yield of ·OH with respect to pH was proposed to account for the degradation kinetics of organic pollutants during sonication.

Table 2 Summary of existing explanations for pH effects on sonochemical degradation of organic pollutants
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Although the proposed explanations are valid in each individual case, they are not necessarily valid in all cases implying further study is required for clear interpretation of pH effects in sonochemical processes, as cavitation bubble dynamics, molecular properties of organic pollutants, and solution matrix all contribute to the complexity of sonochemical responses (e.g., radical quantity, bubble charge, and number of organic molecules accumulated on bubble–water interface). In particular, identifying and elucidating radical reaction pathways in the ultrasonic or combined systems is crucial to determine the radical yield and related degradation variance with respect to pH. In addition, the potential interference of coexisting buffer ions (e.g., \( {\text{H}}_{ 2} {\text{PO}}_{3}^{ - } \), \( {\text{HPO}}_{3}^{ 2- } \), \( {\text{HCO}}_{3}^{ - } \), and \( {\text{CO}}_{3}^{ 2- } \)) to radical quantification needs to be addressed. Further, the charge of cavitation bubbles needs to be verified and quantified with more concreted evidence. Winter et al. (2009) summarized the surface-selective photoelectron spectroscopy results and molecular dynamics simulations and concluded that the air–water interface was more positively charged than the bulk due to the presence of hydronium ion in acidic solution. This conflicts with observations from Watmough et al. (1992) that cavitation bubbles are negatively charged. Although these studies confirmed that the surface charge on bubbles changes with pH, it is controversial whether the charge becomes more positive or negative as pH changes.

In addition, due to the limited knowledge of fluid and bubble dynamics in ultrasonic reactors, it is quite difficult to fully understand the characteristics of ultrasonic systems and combined effects of the thermodynamics and kinetics of target contaminants. In particular, bubbles in the ultrasound field are subjected to high velocity oscillations and translations (Leighton 1994). These phenomena significantly affect the fluid dynamics in the reactor (Wei et al. 2015; Wei and Weavers 2016), resulting in a more complex factor to take into consideration when it comes to prediction of the degradation kinetics using a single physicochemical property of an organic pollutant. In combined systems such as ultrasound-ozone, it is obviously more problematic to correlate the bubble and fluid dynamics to degradation kinetics, especially for the determination of radical reaction pathways and free radical quantity in the presence of ozone or catalytic UV processes.

In summary, it is of particular interests and importance to conduct the following future studies to clarify the pH effects in ultrasonic systems: (1) quantifying free radical yield and concentration distribution, molecule accumulation on bubble water surface, and electrostatic interaction between compounds and bubbles to evaluate the relative contribution of each mechanism; (2) investigating the bubble and fluid dynamics at different pH conditions to reveal their influence to organic pollutant degradation; (3) improving or designing new analytical technique to overcome the low detection limit of free radicals (e.g., ·HO2 −) in the presence of other radicals and species for current instruments; and (4) examining the influence of composition and concentration of water matrices, especially organic matrix, on the sonochemical treatment of wastewater under different pH conditions.