1 State of the art of the environmental behaviour and removal techniques of the 2 endocrine disruptor 3,4-dichloroaniline 3 ANDREA LUCA TASCA* and ASHLEIGH FLETCHER 4 Department of Chemical and Process Engineering, University of Strathclyde, Glasgow, UK 5 Abstract 6 In recent years, the presence of Endocrine Disrupting Chemicals (EDCs) in wastewater 7 discharges from agricultural and industrial sources, 8 soils, has been reported in the literature. 9 environmental exposure to these substances, and the suggestion that humans could also be at [3-5] [2] [1] fresh- and estuarine-waters, as well as Studies of adverse changes in wildlife, linked to 10 similar risk of adverse health effects, 11 and reduce such risks. 3,4-dichloroaniline (3,4-DCA) has been recognized as an EDC, with 12 regards to endocrine disruption data for both wildlife populations and human health. [5] 13 3,4-DCA is present in the environment as a product of the biodegradation of phenylurea and 14 phenylcarbamate pesticides; [6, 7] furthermore, it can be introduced from industrial and municipal 15 wastewater that is insufficiently purified, or via accidental spills. [8-10] Increasing concentrations 16 of 3,4-DCA in soil and water are the result of its high persistence and accumulation, as well as 17 its low biodegradability. 18 when considering the low removal achieved by traditional activated sludge treatments, and the 19 generation of carcinogenic trihalomethanes as a consequence of the chlorine oxidation methods 20 _________________________ 21 *Address correspondence to Andrea Luca Tasca, Department of Chemical and Process 22 Engineering, University of Strathclyde, Glasgow, United Kingdom. 23 E-mail: andrea.tasca@strath.ac.uk [11, 12] have raised concern for urgent action to understand Hence, remediation techniques require in-depth study, especially [13] Fe0/H2O2 systems, photodegradation using doped 24 frequently used in drinking water plants. 25 TiO2, and the use of dielectric barrier discharge reactors, seem to be the most promising 26 techniques for the removal of 3,4-DCA from water. 27 Keywords: 3,4-DCA; water; environment; pollution; adsorption; electrochemistry 28 Introduction 29 Global annual production of 3,4-dichloroaniline (3,4-DCA) was ~42-47 kt prior to 1986 [14] and, 30 although recent years have seen a decrease in scale, EU production was still around 13.5- 31 15.5 kt annum-1 in the period 1996-1998. 32 intermediate in the chemical synthesis of 3,4-dichlorophenylisocyanate, to make phytosanitary 33 products, such as propanil, linuron, diuron, and neburon, [15-17] used to treat crops including rice, 34 potatoes, beans and tobacco. 35 fabrics 36 from direct use of 3,4-DCA, indirect contact is expected via food, primarily fruit and vegetables, 37 [5] 38 acylchloroanilide pesticides present in soils, 39 Furthermore, industrial wastewater [11] may contain 3,4-DCA, mainly from microbial conversion 40 of 3,4-dichloro-1- nitrobenzol within water treatment plants. [5, 28] 41 In this manuscript, we summarize the current knowledge concerning the introduction, 42 movement, and fate of 3,4-DCA in the environment. We discuss the state-of-the-art remediation 43 technologies currently in use, as well as those under development, with reference to their 44 effectiveness for pollutant removal from soil and water systems. 45 Toxicity 46 In humans, EDCs are known to affect male and female reproductive organs, thyroid metabolism, 47 [29-31] [17] [18] and pharmaceuticals. [15] 3,4-DCA, a derivative of aniline, is an It is also employed in the synthesis of azo dyes for polyester [5] Although there is no exposure risk for vulnerable groups as a result of the hydrolysis and biological degradation of phenylurea, phenylcarbamates and [19-27] by field waters and plant enzymes. breast development, cardiovascular and neuro-endocrinology, [32] [28] causing obesity and 48 some cancers, including prostate cancer. [33] Little is known about the mechanisms of action of 49 these substances, nor their physical and chemical diversity, hence, additional research into EDCs 50 is required, especially on the cumulative impact of EDC mixtures, which may be additive or 51 synergistic, 52 effects. 53 the assessment of human health risk and impact. [36] 54 Chloroanilines can negatively affect soil microflora, and the presence of phenylamide herbicide 55 degradation products has been shown to inhibit Nitrosomonas, hence, soil nitrification. [37] 3,4- 56 DCA primarily acts by polar narcosis on aquatic organisms; 57 extremely sensitive with respect to water column exposure, while annelids are greatly affected 58 by exposure to the sediment. 59 3,4-DCA on marine and fresh water animals, as well as unicellular algae, but there are 60 significant chronic effects 61 changes in secondary sex characteristics; [42] while rats have shown significant hormonal effects. 62 [3] 63 proposed, which are 0.2 μg L−1 for freshwater and seawater bodies, and 0.1 mg kg-1 dry weight 64 (0.04 mg kg-1 wet weight) by mass [17] for sediments. 65 Environmental behaviour and fate 66 Chloroanilines are known to diffuse easily within the natural environment, 67 difficult to remediate, hence, their concentration in soils and waters is increasing due to their 68 high persistence, accumulation, 69 By contrast, 3,4-DCA is barely measured in water bodies, 70 concentrations than the parent herbicide diuron. 71 waters, [35] [32, 34] even when individual chemicals are present below the threshold of detectable Consequently, the current lack of knowledge regarding exposure scenarios hinders [17] [40, 41] [38, 39] fish and crustaceans are Tests have demonstrated the relatively low, acute toxicity of with consequences for marine life reproduction, [4] including As a consequence of these impacts, Predicted No-Effect Concentrations (PNECs) have been [16, 46] [11, 12] low biodegradation [45] [43] [5] [1, 14] and are very and low leaching potentials. [44] and it occurs in higher aqueous It has been detected in European surface and its migration is determined by transport and retention mechanisms. [47] 72 Hydrolysis of 3,4-DCA is not considered a mitigating mechanism due to aromatic stabilisation 73 effects, 74 volatilise from water columns. [17, 48] Rather, as the physical data presented in Table 1 show, 3,4- 75 DCA losses from natural waters occur via photochemical degradation, or adsorption on sediment 76 and dissolved humic materials. 77 pathway of 3,4-DCA in environmental waters, 78 spectrum observed at sea level. [4] 79 Table 1. Properties of 3,4-dichloroaniline [48] and, as a result of its relatively low Henry's constant (KH), it is not expected to [49] To this end, photo-transformation is the major degradation [16] maximised at 300 nm, well within the solar Property Value Ref. Molecular formula C6H5Cl2N [15] Molecular structure Appearance Solid at 293 K [15] Molecular weight 162 g mol-1 [15] Molecular size 0.35 nm2 [50] Henry's constant 0.05 Pa m3 mol-1 [15] Solubility in water 580 mg L-1 at 293 K [15] Octanol-water partition coefficient (log Kow) 2.7 (shaken flask method) [5, 16] Estimated surface water half life 18 days [15] 0.11 - 0.17 day-1 [4] 0.06 - 0.14 day-1 [49] Estimated atmospheric half life 9 hours [15] Estimated half-life in soil and sediment 470 - 1500 days [51] Measured rate of loss from outdoor water systems 80 It is known that soil mobility and bioavailability of pollutants, and their degradation products, 81 depend on a combination of adsorption and desorption by soil components. 82 effect of soil generally increases with time, leading to a decrease in pollutant bioavailability and 83 toxicity, due to meteoric leaching. 84 environment accumulates, over time, on the organic fraction of sediments and soils. [15] On the 85 other hand, sorption on dissolved macromolecules and colloidal particles promotes transport in 86 subsurface environments. 87 only a small portion of the former is mineralized, consequently, many chloroanilines persist for 88 years, [54] often immobilized by interaction with humic substances, as mentioned above. [9, 20, 55, 89 56] 90 polymerization, by microbial oxidases and peroxidases, forming stable azo compounds. 91 Surface adsorption, due to van der Waals forces or electrostatic interactions, is often the initial 92 phase of pollutant binding by soil, while stronger bonds may occur over time. [1] Consequently, 93 fulvic and humic acids play a significant role in the binding of xenobiotics, such as 94 chloroanilines, from aqueous media, via functional substituents, including hydroxyl, carboxylic 95 acid, ketone, amino acid, saccharide and aminosaccharide groups. [58, 59] Aniline sorption to soil 96 involves stronger interactions, starting with hydrophobic partitioning and cation exchange, 97 before covalent bonding, due to the contribution of limited energy or availability of sorption 98 sites, occurs. 3,4-DCA adsorbs onto sediment 99 probably, covalent bonds with organic substances. [53] [52] [47] The binding Hence, the majority of 3,4-DCA released into the When chloroanilines are released into the soil with herbicides, Therefore, it is only a small fraction of liberated chloroanilines that undergo dimerization or [49] [47, 57] and soil particles, building stable, most [16, 17] Such initial reversible equilibrium, [60, 100 followed by a slower irreversible mechanism, is well described by a biphasic kinetic model. 101 61] 102 development of remediation strategies. In laboratory experiments, more than 70% of radio 103 labelled 3,4-DCA was found, as stated above, to bind to sediment and suspended matter in a 104 water column; [62] ~80% of radioactivity was removed from the water column within 8 days, and Similar information on the sorption mechanisms of 3,4-DCA is fundamental to the [63] 105 ~99% after 90 days, 106 significant time to reach its final equilibrium. As a consequence of the proposed interaction 107 between 3,4-DCA and organic materials, the interactions between dissolved organic matter from 108 soil, organic contaminants, and other soil components strongly affect the fate of 3,4-DCA in soil 109 and water systems. 110 amendments to a soil sample, which may also introduce dissolved organic matter, can 111 potentially enhance sorption, and decrease leaching, of pollutants, [68, 69] including 3-4-DCA; [49] 112 notably, similar results have been observed for organic matter amendment of inorganic soils. 113 47] 114 A study into the interaction of various humic fractions and the herbicide diuron 115 the main degradation product, 3,4-DCA, was irreversibly bound on humic acids within days of 116 formation. As a result, the risk of pollutant leaching is expected to be low for soils with high 117 humic or fulvic acid contents. The irreversible sorption observed in these systems was described 118 using a Freundlich isotherm model: Cs = K f Cen suggesting an initially quick process, which then plateaus and requires [64-67] As a result, increasing soil organic carbon content, via organic matter [70] [8, showed that (1) 119 where Cs is the concentration of 3,4-DCA sorbed (μg g-1), Ce is the equilibrium solution 120 concentration of 3,4-DCA (μg mL-1) and Kf (mL g-1) expresses the soil sorption capacity. The 121 exponent n is related to the degree of isotherm nonlinearity, and provides an indication of the 122 favourability of the sorption process. [71] 123 The results of Freundlich analysis indicated that humic fractions have a sorptive capacity ten 124 times that of their fulvic counterparts, due to preferential reaction of the amino groups of 3,4- 125 DCA with the carboxyl and carbonyl groups of soil humic fractions, leading to the formation of 126 soil bound residues, [72, 73] as confirmed by high adsorption and small desorption rates constants. 127 [49] Hence, the sorption of 3,4-DCA in soil can be described as a physicochemical process, in 128 which a fraction of the pollutant physically binds to organic and inorganic soil components, 129 while another fraction strongly adsorbs on the organic component. 130 agitated in aqueous solutions of 3,4-DCA, have shown significant pollutant removal within 131 15 minutes of exposure and full equilibration, demonstrating up to 70% removal, after 50 hours. 132 The authors of this previous study reported the presence of two associated processes, firstly a 133 physical reaction, seemingly followed by chemical bond formation between 3,4-DCA and 134 organic matter within the soil, giving biphasic kinetics with rate constants of 4.9 hour-1, for 135 physical accumulation, and 0.03 hour-1, for chemisorption. 136 been observed for different agricultural soils and 3-4-DCA, 137 reached within 48 hours regardless of initial pollutant concentration, up to 16.2 μg mL-1. Again, 138 the data was satisfactorily described by the Freundlich equation, and the amount sorbed was 139 highest for the sandy clay loam soil used in the study (Kf = 52 mL g-1), as a consequence of its 140 higher organic matter content compared to the other soils studied, and the fact that it creates a 141 slightly acidic pH within the sorption system. By contrast the lowest sorption capacities were 142 obtained for calcareous silty clay soil and sand. 143 As initial adsorbate concentration increases, the availability of adsorption sites decreases, as 144 confirmed by n values lower than 1; such sorption behaviour being expressed by L-type 145 isotherms. 146 soil sample studied, show significant probability for 3,4-DCA contamination of ground-water 147 for soils with lower sorption potentials. 148 between the sorption capacity and organic matter content of soil samples, and the fact there is 149 less impact with respect to clay content or a material’s cation exchange capacity. It is also 150 noteworthy that consideration of soil organic content alone is insufficient to determine expected 151 sorption behaviour; for example, a lower level of diuron sorption is expected on clay-rich soils, 152 most likely as a result of a reduction in available binding sites in humic substances, due to the [74] [49] [43] Similarly, soil samples, Similar kinetic performance has [43] with sorption equilibrium Koc values, representing the sorption constant per gram of organic carbon in the [44] This is due to the strong relationship exhibited [70] 153 positive interactions between humic materials and clay. 154 impacts on sorption of 3,4-DCA, as demonstrated by a reduction in Kf values, by ~50%, after 155 liming of aqueous solutions. 156 functionalities within the soils tested, which are normally protonated species under the usually 157 acidic conditions found in such media, becoming neutral species as pH increases. 158 Samples of calcareous soil mixed with aqueous 3,4-DCA, and allowed to equilibrate fully 159 showed high Kf values, which indicates that adsorption is concentration dependent, 160 validating the assumption that adsorption on such solid media occurs primarily via hydrophobic 161 interactions due to the neutralisation of aniline functional groups at high soil pH. 162 Additional confirmation is provided by consideration of the physical properties of 3,4-DCA; Kow 163 is known to be high, 164 (Koc = Kd/[organic content] × 100, where Kd is a measure of the distribution ratio of organic 165 molecules between the sorbed phase and solution), this all suggests that 3,4-DCA has a low 166 potential for groundwater contamination in calcareous soils with high organic matter content. 167 Dissolved organic matter has been proven to compete with organic pollutants for the sorption 168 sites available on soil surfaces, [80, 81] as well as in the building of stable bonds between pollutant 169 species and soil; 170 thereby reducing sorption and increasing their mobility. 171 materials applied in soils may actually be adsorbed to soil surfaces, increasing sorption of 172 hydrophobic organic compounds, 173 dissolved organic carbon extracts, derived from both a commercial peat and high-purity tannic 174 acid, to a soil sample showed significant impact on the sorption of 3,4-DCA. 175 coefficient, defined as Xdoc/Cdoc (where Xdoc is mg L-1 of dissolved organic carbon and Cdoc is 176 the corresponding equilibrium concentration), is consistently higher than the Kd values obtained 177 for undoctored soil, thereby confirming the influence of dissolved organic carbon on 3,4-DCA [82] [9, 79] [43] Soil pH may be another factor that Such a trend probably results from the presence of aniline [61, 75, 76] [77] [47] thereby [73, 78] its solubility on water is low, combined with a Koc of 338.6 L kg−1 this competition enhances the apparent solubility of organic pollutants, [84] [67, 83] In contrast, dissolved organic especially when tannic acid is added. [8] Inclusion of [47] The Kd,DOC 178 sorption. Further confirmation is provided by an increase in 3,4-DCA sorption in the presence 179 of an environmental matrix, including inorganic ions and organic matter. 180 sedimentary material, hosted in a sediment extract media, was higher than for pure or run-off 181 waters, likely facilitated by previously sorbed dissolved organic content on the sediment surface, 182 present as a result of physical interactions between the two species. [73] 183 Remediation technologies 184 Remediation technologies developed for the destruction of chloroanilines present in wastewaters 185 can be classified as physicochemical, chemical (oxidation) or biological, and are discussed in 186 detail below. In essence, physicochemical methods utilise either adsorption, ion exchange, 187 electrolysis or photodegradation, chemical processes require a chemical reaction with a selected 188 additive, while biological degradation involves the action of aerobic or anaerobic 189 microorganisms. 190 Bioremediation 191 Bacteria 192 3,4-DCA is not readily biodegradable; [17] the process being particularly slow in aqueous media. 193 Incubation of pond water, and pond water containing sewage sludge inoculum, in a darkened 194 environment showed that, after a period of 2 weeks, 97% of 3,4-DCA was recovered from the 195 former sample and marginally less (94%) was recovered for the sample containing sewage, 196 indicating little biodegradation of the pollutant. 197 over the same time period, for an OECD 301 C test 198 4 weeks for an OECD 301 D test on activated sludge, and less than 5% degradation was 199 observed after 29 days for an OECD 303 A test, 200 Researchers also observed no discernible removal of 3,4-DCA from contaminated North Sea [86] [85] Sorption on Similarly, no biodegradation was reported, [87] [87] with activated sludge, [88] nor after again using activated sludge. [17] 201 water samples, [89] while only primary degradation occurred after one month using river water as 202 an inoculum. [90] 203 Usually, xenobiotics need to be in an aqueous phase in order to allow bacterial degradation to 204 occur. The ‘bioaccessible fraction’ of a pollutant is given by the sum of its concentration in pure 205 water, known as the ‘bioavailable fraction’, plus the ‘potentially available fraction’, which is 206 the material reversibly sorbed on any material surfaces. The addition of fulvic and humic acids 207 to inoculated soils was seen to decrease the rate of diuron degradation, reducing bioavailability 208 but not bioaccessibility, hence, lengthening the treatment time required; so it was only after 209 32 days that all diuron was degraded to 3,4-DCA. 210 indicator than bioavailability of the long-term influence of humic substances on diuron 211 degradation. The mineralisation rate of 3,4-DCA in soils is low, and it decreases as pollutant 212 concentration increases; [15] only 3.9-11.9% mineralisation of 1 mg kg-1 radio labelled 3,4- DCA 213 was recorded after 16 weeks within various soil types. [51] Degradation of 50% of 3,4-DCA was 214 observed in soil slurries with indigenous soil populations, and this was only marginally 215 influenced by the addition of buffer, mineral salts and acetate. [91] In non-acclimated sediments, 216 dechlorination of applied 3,4-DCA started after 20 days, with anaerobic conversion to 3- 217 chloroaniline (44%) and 4-chloroaniline (33%) within two months; these metabolites were not 218 further degraded. [92] 219 The microbial strains Pseudomonas acidovorans 220 use chlorinated anilines as a sole source of carbon and energy ; the latter also being capable of 221 growing on 3,4-DCA. 222 slurries enhanced pollutant mineralization, leading to complete elimination of chloride after 223 10 days. [91] Up to 250 mg L-1 of the pollutant and its intermediates were anaerobically degraded, 224 in under 7 days, by a strain of Pseudomonas fluorescens. Without added glucose and nitrogen [95] [93] [70] Hence, bioaccessibility is a better and Pseudomonas diminuta [94] are able to Addition of Pseudomonas acidovorans to 3,4-DCA enriched soil 225 sources, degradation was slower, with 40% of toxicant removal in the first 15 days, at an initial 226 concentration of 75 mg L-1. 227 50 μg mL-1 of 3,4-DCA within its growth process, increasing the ratio of degradation in water 228 samples from natural water reservoirs 229 reversibly sorbed pollutant was observed within the first 12 days. It is also notable that an 230 appreciable decrease in the irreversibly sorbed fraction occurred within the first 5 days. [98] 231 Microorganisms, from Cuban soils, were grown in two culture media, using 3,4-DCA, firstly as 232 the sole source of carbon and secondarily as the sole source of carbon and nitrogen. 233 pollutant was completely consumed within 3 weeks using Pseudomonas, Arthrobacter, 234 Aspergillus, Penicillium, and Fusarium, isolated in the first medium, while Bacillus, 235 Arthrobacter, Cunninghamella, Trichoderma, and Fusarium were isolated in the second system, 236 demonstrating myriad bacterial growth from 3,4-DCA as a feedstock. The biodegradation 237 pathway of 3,4-DCA, and other substituted anilines, involves conversion, by oxygenase, to the 238 corresponding catechol, which is then metabolised via an ortho-cleavage pathway. 239 modes of dioxygenation have been determined, utilising degrading bacteria obtained by genetic 240 exchange between two strains of Pseudomonas, and leading to the formation of 3- and 4- 241 chlorocatechol from 3-chloroaniline. In contrast, only 4-chlorocatechol was generated from 242 dioxygenation of 4-chloroaniline. [94] When bacterial strains of Pseudomonas acidovorans were 243 used for the degradation of 3-chloroaniline and 4-choloranaline, the rate-limiting degradation 244 steps were found to be the first attack of the substrate, and conversion to chlorocatechols. [93] 245 Degradation of 3,4-DCA by Pseudomonas sp. showed that catechol 2,3-dioxygenase is integral 246 to process efficiency; 247 chlorocatechol was found to be 60, 27 and 13%, respectively, of the activity toward catechol 248 2,3-dioxygenase. Further tests confirmed catechol 2,3‐dioxygenase activity using a strain of [95] [96] Pseudomonas diminuta was proven to dechlorinate up to [97] and in fish ponds, where significant degradation of [99] [100] The Two activity toward 4-methylcatechol, 3-methylcatechol and 4- 249 Pseudomonas fluorescens; the presence of 3‐chloro‐4‐hydroxyaniline as a metabolite suggesting 250 a pathway that includes dehalogenation and hydroxylation of the aromatic ring, followed by ring 251 cleavage, by catechol 2,3‐dioxygenase. [96] 252 Recent work has shown Micrococcus sp. to degrade 96% of diuron within 30 hours of 253 incubation, at a concentration of 250 ppm, and with the addition of non-ionic detergent (0.01%). 254 [101] 255 hydrolysis step, leading to the accumulation of 3,4-DCA, which undergoes conversion to 4,5- 256 dichlorobenzene-1,2-diol, and further intermediates, within 24 hours of test commencement 257 (Fig. 1). 258 Arthrobacter sp. and Achromobacter sp., with CO2 as the only final product. [102] The authors proposed a mechanism whereby a methyl group is removed, followed by a Diuron mineralization has also been confirmed by the metabolic cooperation of 259 260 Figure 1. Degradation pathway of diuron by Micrococcus sp, confirmed by FTIR spectra and 261 HPLC [101] 262 Strains of Aquaspirillum itersonii, Aquaspirillum sp. and Paracoccus denitrificans were shown 263 to successfully use 3,4-DCA as the only source of carbon and nitrogen for growth; [103] the latter 264 was able to metabolize the pollutant at concentrations up to 150 mg L-1, through oxidation to o- 265 diphenol, intradiol cleavage of 4,5- dichloropyrocatechol, and further stages of preparatory 266 metabolism associated with dehalogenation. A study of microorganisms isolated from Cuban 267 soils, treated with propanide, showed a Paracoccus denitrificans strain to be most efficacious for 268 3,4-DCA 269 following formation of 4,5- dichloropyrocatechol, allows full decomposition. Successful 270 adsorptive bioremediation was demonstrated with the introduction of activated carbons and 271 bacteria to polluted soils. 272 concentration in the soil solution below the toxicity threshold for the bacteria, as shown by a 273 study of three types of activated carbon saturated with 3,4-DCA and placed in a mineral medium 274 with a Paracoccus denitrificans strain. 275 available, at a limited concentration, for the bacteria population to process. Varied degradation 276 rates, from 2 to 10 weeks, suggest that facile desorption and more rapid decomposition are 277 linked to a reduced micropore volume. This can be ascribed to the fact that Paracoccus 278 denitrificans is only able to penetrate into the macroporous structure, due to size exclusion 279 effects, thus pollutant desorption from the smaller carbon pores is necessary for the bacteria to 280 process any 3,4-DCA sorbed therein; this bacterium seems able to accelerate desorption by 281 acidifying the medium or via excretion of surface-active substances. 282 Biological studies have also shown microbial consortia to be successful in the removal of 283 propanil and 3,4-DCA from repeated batch suspended cell cultures, 284 reactors for agricultural wastewater treatment. 285 tezontle was used as a support for a biofilm in a continuous process able to degrade propanil and 286 metabolic intermediates at rates of up to 24.9 mg L-1 h-1, without the need for co-substrates. [50] degradation. The cleavage of the aromatic ring via the ortho- or meta-pathways, [104] Activated carbon acts as buffer, which keeps the pollutant [50] The pollutant was reversibly sorbed, therefore [106] [105] as well as in biofilm In the latter case, the porous volcanic stone 287 Pseudomonas sp., Acinetobacter calcoaceticus, Rhodococcus sp., Xanthomonas sp. and Kokuria 288 strains can also grow individually in 3,4-DCA, while other strains found in the biofilm, not able 289 to degrade propanil metabolites directly, are probably involved in the metabolization of 290 herbicide adjuvants or in the maintenance of biofilm integrity. Resultantly, the removal of 291 chloroanilines from sewage treatment plant streams could be improved by promoting the growth 292 of indigenous bacterial communities, and through the introduction of adapted laboratory strains. 293 The addition of readily degradable aniline and non-toxic haloaromatics may, respectively, 294 improve the breakdown of chloroanilines and the chlorocatechol potential. [94] 295 Uptake by fungi and cultivated plants 296 When free chloroanilines are released as herbicide metabolites, they can be incorporated in the 297 plant’s ‘insoluble’ residue fraction; degradation experiments have proposed lignin as a primary 298 binding site. 299 downstream of conventional rice fields, in the region of Camargue, [19] showed the concentration 300 of 3,4-DCA measured downstream of the rice plantations to be approximately half that in the 301 Corbicula caged upstream, suggesting a potential, partial bioaccumulation of 3,4-DCA in rice 302 plants. Tomato plants, oat, barley and wheat, grown in nutrient solutions with 4-chloroanaline 303 and 3,4-DCA showed that 90-95% of the chloroanilines incorporated were found in the roots, 304 with uptake proportional to the amount of chemical applied. In contrast, distribution of the same 305 chemicals in carrots was approximately equally divided between the roots and the upper part of 306 the vegetables. 307 assumption of 3,4-DCA contained in certain foods. 308 Some lignin degraders are able to metabolize chloroanilines and their lignin conjugates; 309 experiments have shown that chloroanilines appear to be bioavailable to the white rot fungus 310 Phanerochaete chrysosporium once they were mineralized as lignin. [107, 108] [109] Immersion of the bivalve Corbicula fluminea in cages both upstream and These results suggest a potential risk of chronic toxicity due to the [110] More than 50% of 311 available [ring-U- 14C] -3,4-DCA was shown to be mineralized after 33 days of sample 312 incubation and free 3,4-DCA was deemed a superior substrate for mineralization than free 4- 313 chloroanaline. Different metabolites were formed, but chloroanilines were not detected, neither 314 were their azo or azoxy derivatives. Hence, lignin incorporation and fungal oxidation can lead 315 to the complete removal of 3,4-DCA from the environment; however, fungi is also known to 316 adsorb less chloro-substituted anilines per biomass unit than bacteria, 317 pathway of white-rot fungi could lead to the formation of toxic tetrachloroazobenzenes. [112] 318 Adsorption and ligand exchange 319 Adsorption technologies offer effective removal of many organic pollutants from aqueous 320 media, 321 wastewaters. Batch adsorption experiments, conducted with an acid activated halloysite, using 322 aqueous solutions at pH ~5, 323 kinetic model: [114] [113] [111] and the degradation and various adsorbents have been studied for removal of chloroanilines from 𝑡 1 1 = + 2 𝑞𝑡 (𝑘2 𝑞𝑒 ) 𝑞𝑒 𝑡 [12] gave experimental data that followed a pseudo-second order (2) 324 where: qt is the amount of chloroaniline adsorbed (mg g−1) at time t (s), k2 is the rate constant of 325 pseudo-second order adsorption (g mg−1 min−1) and qe is the amount of chloroaniline adsorbed at 326 equilibrium (mg g−1). A ‘Weber–Morris’ plot of qt versus t0.5 confirmed that chloroaniline 327 removal occurred, first by fast diffusion of 3,4-DCA to the surface of the clay mineral, over the 328 first 180 minutes, before continuing as slower interparticle diffusion. The adsorption capacity of 329 halloysite was found to be lower for 3,4-DCA than for 3-dichloroaniline and 4-dichloroanaline. 330 Equilibrium isotherm data was effectively described by the semi-empirical Langmuir equation: 𝐶𝑒 1 𝐶𝑒 = + 𝑞𝑒 (𝐾𝐿 𝑞𝑚 ) 𝑞𝑚 (3) 331 where: qm is the monolayer adsorption capacity (mg g-1), qe is the sorption uptake at equilibrium 332 (mg g-1), Ce is the equilibrium solution concentration (mg L-1) and KL is a coefficient related to 333 the affinity between the adsorbent and the adsorbate (L g-1). Similar adsorption experiments 334 have been performed with kaolinite (KGa-1) and montmorillonite (SWy-1), using standard 335 solutions of 3,4-DCA and other chloroanilines at pH ~5 and ~9, respectively; 336 pH conditions were chosen so as to produce neutral organic pollutant species. 337 evaluation indicated an initial, rapid surge in chloroaniline removal, with equilibrium achieved 338 in under 4 days. Langmuir and Freundlich equations both adequately described the data; 339 however, the Langmuir plot showed a marginally better fit for montmorillonite, hence 340 adsorption on this clay is likely to decrease as surface sorption sites are saturated. 341 chloroanilines studied would be mainly sorbed on the mineral surface of kaolinite, while the 342 structure of montmorillonite allows swelling via sorption in the interlayer, 343 studying the dehydrated clay; 344 planes (d001) collapsed from 11 to 9.7 Å. Further evidence was provided by X-ray spectra of the 345 montmorillonite/3,4-DCA system. 346 desorption measurements, where kaolinite was shown to retain the pollutant, while 347 montmorillonite showed a higher level of desorption when reversing the sorption process at an 348 earlier point in the isotherm. [75] 349 While clay materials offer surface sites for adsorption, the main surface area of another well- 350 known family of sorbents, i.e. activated carbons, is ascribed to microporous character; surface 351 hydrophobicity determines the sorptive capacity of many organic molecules, which have 352 molecular sizes small enough to penetrate into the micropores. Highly microporous activated 353 carbons, obtained from coal and peat, were confirmed as better sorbents than mesoporous 354 carbons, obtained from raw plant materials. Increased iron content and other ash elements may [1] [115] [1, 75] the specific Kinetic [75] The as confirmed by post heat-treatment, the distance between equivalent atomic This contrast in sorption mechanism is reflected in 355 positively influence the maximum uptake, by enhancing chemisorption of active organic 356 compounds, such as 3,4-DCA. [50] 357 Studies of aqueous solutions of various pesticides agitated with 10 mg L−1 of powdered activated 358 carbon (surface area ~1000 m2 g-1; particle size 40 µm) for 5 minutes, showed 70% removal 359 efficiency for 3,4-DCA (initial concentration: 658 ng L-1), while complete removal was achieved 360 via preoxidation with ozone. [103] Fitting of room temperature sorption isotherms, obtained using 361 suspensions of activated carbon in aqueous solutions of 3,4-DCA, 362 isotherm model (Equation 3), showed monolayer filling of the sorptive surface with ‘L-type’ 363 isotherms, indicating strong interactions between the adsorbate and adsorbent. [116] 364 Sugar beet pulp, corncob, corncob char, perlite, vermiculite and sand have recently been studied 365 for sorption of 3,4-DCA from aqueous solutions at pH 4.8. 366 field conditions, so it is essential to understand their sorption behaviour in different matrices. 367 However, liquid matrices containing cations and organic matter, except for corncob, showed no 368 significant differences in maximum uptake of 3,4-DCA from pure water and run-off waters. 369 The mechanism of sorption was satisfactorily described by the Freundlich isotherm model, with 370 the highest sorption uptakes obtained at 99% removal from water for corncob char, and, 86% 371 removal for sand. 372 resistance to mechanical abrasion. 373 adsorption capacity of low cost materials in field conditions, but it is encouraging that sorbents 374 such as sand could present appreciable capacities capable of removing > 50% of 3,4-DCA, even 375 from sediment extract media (Table 2). 376 Table 2. Surface area, pore size and adsorption data of different material tested for the removal 377 of 3,4-DCA from water. The parameters qm and b refer to the Langmuir model (Equation 3) [85] [50] using the Langmuir These sorbents may be used in Of the sorbents studied, it is also known that vermiculite has a good [117] Further investigations would be required to verify the Adsorbent Surface Pore area volume Equilibrium qm b (mg g-1) (mL mg-1) time Stirring method (m2 g-1) (cm3 g-1) Halloysite [12] 76.6 0.039 0.078 2.726 >3 Rotary stirrer Kaolinite [75] - - 0.311 9 > 96 Electromagneti (h) c Montmorillonit - - 0.077 23 > 96 e [75] Activated Electromagneti c 963 0.55 583 - 48 - 1028 0.53 480 - 0.5 - 410 0.5 364 - 0.5 - carbon AG [50] Activated carbon SKT [50] Activated carbon RS [50] 378 Within ligand exchange processes, polymeric chelating resins are able to selectively remove 379 target contaminants; however, eluate recycle, regeneration of depleted adsorbent, and the high 380 cost of transition metals, used as ligand complexing ions, are still significant obstacles towards 381 commercial application of such processes in wastewater treatment. Currently, there are no 382 studies on the application of ligand exchange processes for 3,4-DCA removal, but this would 383 seem possible in light of the recovery of aromatic amines from water, at low concentration, as 384 demonstrated for chelating resin-bound cobalt ions. 385 CDAE-sporopollenin resin was also tested for adsorption of chlorinated anilines, found to be 386 described by a Langmuir model for 2-chloroaniline, 4-chloroaniline and 2,5-dichloroaniline. 387 The study showed similar values for the maximum adsorption capacity (qm) for binding of 2- [118] A mini-column apparatus with Co(II)- [6] 388 chloroanilines and 4-chloroanilines, which were consistently lower than those for 3- 389 dichloroaniline and 2,5-dichloroaniline onto the Co2+ matrix, suggesting that both electrical 390 forces and steric hindrance are involved in the sorption process. This conclusion is supported by 391 consideration of the inductive effects of ortho-Cl and para-Cl atoms, as well as the nature of 392 these ligands, which contain charged groups and may offer steric hindrance. Moreover, steric 393 hindrance around the amino nitrogen weakens binding to metal ions, causing faster migration of 394 the aforementioned pollutants. 395 a Freundlich model, possibly as a result of a more complex type of binding than the independent 396 and univalent binding described by a Langmuir model. 397 Chlorination, ozonisation, chemical precipitation and Fe0/H2O2 systems 398 A common sequence of operations adopted in many drinking water plants is that of (i) 399 preoxidation, (ii) adsorption, and (iii) coagulation. Preoxidation of an aqueous sample with a 400 concentration of 658 ng L-1 of 3,4-DCA, performed using sodium hypochlorite, demonstrated 401 100% pollutant removal; 402 trihalomethanes by-product formation. 403 species and ozonolysis of a sample, again with a concentration of 658 ng L-1 of 3,4-DCA, 404 showed 85% pollutant removal; however, the subsequent coagulation and flocculation steps 405 were found to be ineffective, while further adsorption treatment, with activated carbons, led to 406 complete pollutant removal. [13] 407 The oxidizing potential of Fe0 towards different organic compounds is well known; 408 H2O2 systems can be used to reduce levels of diuron, and other pesticides, in polluted 409 environments, as well as agricultural waste. Fe0 promotes the reduction of H2O2 to hydroxyl 410 radicals, generating Fe2+, which, in turn, also produces hydroxyl radicals via further H2O2 411 reduction. A 10 mg L-1 diuron solution, also containing 2 mmol L-1 of H2O2 and H3PO4, was [13] [119] The adsorption of 3-chloroaniline was better represented by however, such treatments have an associated risk of carcinogenic Hence, researchers have considered other oxidative [120] Fe0/ 412 flowed through a glass tube packed with 2 g of iron wool, and showed that pH strongly affected 413 the degradation process, allowing process optimisation. At pH 2.5 more than 99.9% of the 414 pesticide was removed after ten minutes, with only 1 µg L-1 of 3,4-DCA found remaining in the 415 effluent. [121] 416 Electrochemical and electrohydraulic methods 417 An electrohydraulic discharge (EHD) method for the oxidative degradation of 3,4-DCA was 418 tested by exposing wastewaters to pulsed electrical discharges generated via submerged 419 electrodes. 420 generating a shockwave as it expands against the water. The degradation rate is expressed by: [122] UV radiation is produced by a plasma channel created by EHD, thereby, 𝑑𝐶 = −𝑘1 𝐶𝑖 − 𝑘0 𝑑𝑁 (4) 421 where dC/dN is the change in concentration per discharge, Ci is the initial substrate 422 concentration, k0 is the zero-order term (an expression of direct photolysis) and k1 is the first- 423 order term related to oxidation in the plasma channel region. 424 As part of an Advanced Oxidation Process (AOP), photocatalysis can be employed (i) for water 425 treatment in slurry reactors, where an additional step is required for the separation of any 426 suspended catalysts, or (ii) into reactors, where the catalysts are immobilized on adsorbents or 427 on membranes. Using sols of vanadium pentoxide and cerium oxide, added to a titanium 428 dioxide sol, allowed preparation of Ti–V and Ti–Ce catalysts, respectively; 0.1 g L−1 of each 429 powdered catalyst was added to agitated aqueous solutions of 3,4-DCA, irradiated in an annular 430 reactor at 140 mW cm−2, and the Ti–V catalyst gave a higher degradation than Ti-Ce, due to its 431 band gap energy (which is more towards the visible region) and smaller particle size. The 432 kinetic plot suggested bi-phasic kinetics, with a sharp increase in rate after 45 minutes; further 433 bench scale reactor experiments showed 85% degradation in 106 minutes [11] and Figure 2 shows 434 the intermediate species formed. 435 436 Figure 2. Intermediates formed during photo-degradation of 3,4-DCA using Ti-V (sol) catalyst, 437 confirmed by HPLC and GC-MS analysis [11] 438 Biphasic kinetics were also observed in the photocatalytic degradation of 3,4-DCA using Ti–N 439 and Ti-S catalysts; 440 band-gap energy. Higher specific surface area increases the degree of contact of the pollutant 441 with the catalyst surface, while smaller particle size means a reduced degradation time, due to 442 shorter distances between the charge carrier and the surface, where the reaction occurs. 3,4- 443 DCA was fully degraded using a Ti–N catalyst in 120 minutes with optimal degradation 444 obtained at pH = 6. At higher catalyst dosages (> 0.1 g L−1) the reaction rate starts to decrease, 445 due to the deactivation of molecules that collide with ground state molecules, and subsequent 446 agglomeration of catalyst particles. Meanwhile, at 3,4-DCA concentrations > 10 mg L-1, the 447 degradation efficiency starts to decrease, as the number of collisions between the 3,4-DCA [28] the former showing a higher surface area, and lower particle size and 448 molecules increases, while there is a reduction in collisions between pollutant molecules and 449 OH 450 Degradation of aqueous 3,4-DCA was recently performed using a dielectric barrier discharge 451 (DBD) plasma reactor; [124] where generated ozone interacts with the pollutant directly or by the 452 generation of hydroxyl radicals, which results in a faster rate of reaction than for molecular 453 ozone alone. 454 created by two aluminium electrodes, and the degradation process is well described by pseudo- 455 first order kinetics, with higher efficiencies reached either under acidic conditions, increasing 456 the input power or by adding Fe2+ or Fe3+. The main pathways involved were deamination, 457 hydroxylation, dechlorination, and hydroxylation and oxidation, followed by the generation of 458 organic acids, via aromatic ring opening. Mineralization into CO2 and H2O was only partially 459 achieved, as confirmed by the lower rate of removal for total organic carbon than for DCA, and 460 by degradation intermediates identified using GC-MS analysis. Solution pH decreased during 461 DCA degradation, as observed in previous work, [124] where aqueous 3,4-DCA was degraded by 462 a wire-cylinder DBD reactor, with an efficiency that was observed to increase when the input 463 power was increased to 90 W, but decreased at powers above that; aqueous ozone concentration 464 was seen to follow the same trend. It was also noted that, similar to other systems, the process 465 was pH dependent with a lower degradation rate obtained under neutral conditions than at acidic 466 or basic levels. 467 Conclusions 468 Myriad technologies have been developed for the degradation of pesticides; however, more 469 attention has to be mainly focussed on the fate of metabolites. 3,4-DCA is a degradation 470 product generally more toxic than its parent substances; if it is covalently bound to humic radicals. [123] [125] The process involves the flow of a water film through the discharge zone 471 substances within soil, the risk of groundwater contamination via leaching is low, but it also has 472 an extremely low rate of mineralization, [70] making it a significant environmental issue. 473 Lignin incorporation and fungal oxidation are able to effect complete removal of 3,4-DCA from 474 the environment, 475 bioreactors, as well as for in-situ bioremediation techniques. 476 successfully demonstrated, introducing activated carbons and bacteria to polluted soils, 477 the biological removal of 3,4-DCA from sewage can be enhanced by the growth of indigenous 478 communities, and through the introduction of adapted laboratory strains; 479 technical and economic feasibility of such processes also need to be considered. 480 investigations are required to confirm the adsorption capacity of promising low cost materials 481 such as activated carbons, corncob char and sand, especially within field conditions. 482 possibility of recovering aromatic amines from low concentration water streams has been 483 already demonstrated using chelating resin-bound cobalt ions, [118] as well as for Co(II)-CDAE- 484 sporopollenin resin, 485 technologies to 3,4-DCA removal. 486 Fe0/H2O2 systems could be developed for the degradation of 3,4-DCA in agricultural soils and 487 waste, 488 degradation kinetics are available for the scale up of reactors. 489 combined with good chemical stability and high natural abundance 490 potentially suitable for the environmentally friendly removal of 3,4-DCA from water. Effective 491 degradation is also observed for the use of dielectric barrier discharge reactors, but further 492 studies are required to reach complete mineralization of 3,4-DCA. 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Applied Catalysis A: General. 2008, 340(2), 153-161. 847 Figures and Tables 848 Figures and tables are listed below. 849 Table 1. Properties of 3,4-dichloroaniline Property Value Ref. Molecular formula C6H5Cl2N [15] Molecular structure Appearance Solid at 293 K [15] Molecular weight 162 g mol-1 [15] Molecular size 0.35 nm2 [50] Henry's constant 0.05 Pa m3 mol-1 [15] Solubility in water 580 mg L-1 at 293 K [15] Octanol-water partition coefficient (log Kow) 2.7 (shaken flask method) [5, 16] Estimated surface water half life 18 days [15] 0.11 - 0.17 day-1 [4] 0.06 - 0.14 day-1 [49] Estimated atmospheric half life 9 hours [15] Estimated half-life in soil and sediment 470 - 1500 days [51] Measured rate of loss from outdoor water systems 850 851 852 Figure 1. Degradation pathway of diuron by Micrococcus sp, confirmed by FTIR spectra and 853 HPLC [101] 854 Table 2. Surface area, pore size and adsorption data of different material tested for the removal 855 of 3,4-DCA from water. The parameters qm and b refer to the Langmuir model (Equation 3) Adsorbent Surface Pore area volume Equilibrium qm b (mg g-1) (mL mg-1) time Stirring method (m2 g-1) (cm3 g-1) Halloysite [12] 76.6 0.039 0.078 2.726 >3 Rotary stirrer Kaolinite [75] - - 0.311 9 > 96 Electromagneti (h) c Montmorillonit - - 0.077 23 > 96 e [75] Activated Electromagneti c 963 0.55 583 - 48 - carbon AG [50] Activated 1028 0.53 480 - 0.5 - 410 0.5 364 - 0.5 - carbon SKT [50] Activated carbon RS [50] 856 857 858 859 Figure 2. Intermediates formed during photo-degradation of 3,4-DCA using Ti-V (sol) catalyst, 860 confirmed by HPLC and GC-MS analysis [11] 861