Local applications but global implications: Can pesticides drive microorganisms to develop antimicrobial resistance?

Science of the Total Environment 654 (2019) 177–189

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Review

Local applications but global implications: Can pesticides drive microorganisms to develop antimicrobial resistance? Balasubramanian Ramakrishnan a, Kadiyala Venkateswarlu b, Nambrattil Sethunathan c, Mallavarapu Megharaj d,⁎ a

Division of Microbiology, ICAR-Indian Agricultural Research Institute, New Delhi 110012, India Formerly Department of Microbiology, Sri Krishnadevaraya University, Anantapur 515055, India Flat No. 103, Ushodaya Apartments, Sri Venkateswara Officers Colony, Ramakrishnapuram, Secunderabad 500056, India d Global Centre for Environmental Remediation (GCER) and Cooperative Research Centre for Contamination Assessment and Remediation of the Environment (CRC CARE), University of Newcastle, ATC Building, Callaghan, NSW 2308, Australia b c

H I G H L I G H T S • Pesticides are stored inappropriately, and applied often indiscriminately. • Microbes employ several mechanisms to tolerate, resist and degrade pesticides. • Microbial responses to pesticides may include antimicrobial resistance. • Pesticide-degraders with antimicrobial resistance are of global concern.

G R A P H I C A L

A B S T R A C T

Spread of antimicrobial resistant genes

Intrinsic resistance

Acquired (mutational) resistance

Pesticide-degrading and antimicrobial resistant microorganisms

Microbiome

a r t i c l e

i n f o

Article history: Received 10 August 2018 Received in revised form 2 November 2018 Accepted 3 November 2018 Available online 06 November 2018 Editor: Yolanda Picó Keywords: Pesticides Bacterial degradation Antimicrobial resistance Chemotaxis Efflux pumps ‘One health’

a b s t r a c t Pesticides are an important agricultural input, and the introduction of new active ingredients with increased efficiencies drives their higher production and consumption worldwide. Inappropriate application and storage of these chemicals often contaminate plant tissues, air, water, or soil environments. The presence of pesticides can lead to developing tolerance, resistance or persistence and even the capabilities to degrade them by the microbiomes of theses environments. The pesticide-degrading microorganisms gain and employ several mechanisms for attraction (chemotaxis), membrane transport systems, efflux pumps, enzymes and genetical make-up with plasmid and chromosome encoded catabolic genes for degradation. Even the evolution and the mechanisms of inheritance for pesticide-degradation as a functional trait in several microorganisms are beginning to be understood. Because of the commonalities in the microbial responses of sensing and uptake, and adaptation due to the selection pressures of pesticides and antimicrobial substances including antibiotics, the pesticide-degraders have higher chances of possessing antimicrobial resistance as a surplus functional trait. This review critically examines the probabilities of pesticide contamination of soil and foliage, the knowledge gaps in the regulation and storage of pesticide chemicals, and the human implications of pesticidedegrading microorganisms with antimicrobial resistance in the global strategy of ‘One Health’. © 2018 Elsevier B.V. All rights reserved.

⁎ Corresponding author at: Global Centre for Environmental Remediation (GCER), Faculty of Science, The University of Newcastle, ATC Building, University Drive, Callaghan, NSW 2308, Australia. E-mail address: [email protected] (M. Megharaj).

https://doi.org/10.1016/j.scitotenv.2018.11.041 0048-9697/© 2018 Elsevier B.V. All rights reserved.

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Contents 1. 2.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pesticides in agriculture and public health: the regulatory stringency . . . . . . . . . . . 2.1. Extent of soil and foliar applications of pesticides . . . . . . . . . . . . . . . . 2.2. Obsolete pesticides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3. Microbial degradation of pesticides: a means of cleansing the soil environment . . . . . . 3.1. Chemotaxis: the foremost step for bacterial degradation of pollutants . . . . . . . 3.2. Membrane transport systems: the basis for xenobiotics uptake . . . . . . . . . . 3.3. Efflux pumps: role in antimicrobial resistance . . . . . . . . . . . . . . . . . . 3.4. Plasmid and chromosome encoded catabolic genes: role in pesticide biodegradation 3.5. Enzymes for microbe-mediated pesticide degradation . . . . . . . . . . . . . . 3.6. Evolution and mechanisms of inheritance for microbial degradation . . . . . . . . 4. Blind spots to hotspots: antimicrobial resistance of pesticide-degrading microorganisms . . 5. ‘One Health’: human implications of antimicrobial resistance in pesticide degraders . . . . 6. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1. Introduction Pesticides have been used to control pests and diseases since the beginning of agriculture in the human societies. The global expenditures on the natural and synthetic pesticidal chemical substances are rising steeply to $56 billion with the usage of about 3.5 billion kg of active ingredient yr−1 (Pretty and Bharucha, 2015; Atwood and PaisleyJones, 2017). The cost of research and development for a new active ingredient of pesticides is also huge, more than $286 million with an average time of 11 years from the discovery to market launch (McDougall, 2016). The introduction of new active ingredients with heterocyclic scaffolds, halogen substituents, and enantiomerically or diastereomerically enriched chiral forms has brought the application rates as low as 10 g ha−1 (Lamberth et al., 2013). Though these efforts aim at increasing the efficiencies, this strategy may become counterproductive and follow the consequences of the Jevon's paradox of higher production and consumption of pesticides to adversely influence the ecological economics (Alcott, 2005). Nearly 90% of the total pesticide usage is in the agricultural sector and the rest is used in the nonagricultural sectors that include industry, commercial, and government, and home and garden (Atwood and Paisley-Jones, 2017). Hence, in the present scenario, the food produced for every single meal for just one person uses about 0.3 g of pesticides (Cribb, 2017). Unfortunately, the amount of applied pesticides reaching target pests is b0.1% and the reminder contaminates all other resources. Hence, the contamination of soil, water and air by N99% of the dose applied has serious impacts on public health and beneficial biota (Pimentel, 1995; Schwarzenbach et al., 2006; Guillette Jr. and Iguchi, 2012). Inappropriate application of pesticides (overuse or misuse) or even losses due to spray drift or wash-off from treated foliage or seeds contribute to the soil, water or air contamination (Hildebrandt et al., 2009; Luo and Zhang, 2010). These pesticide chemical substances also get degraded in the environments. The physico-chemical properties of pesticides such as water solubility, lipophilicity, volatility, ionizability, and polarizability, besides the formulation, concentration, method, time, and frequency of application, influence their ‘bioavailability’ and degradation, and the reminder as bound residues (Alexander, 1982; Gevao et al., 2000; Semple et al., 2004; Arias-Estevez et al., 2008). The soil bound pesticide residues (SBPR) could be one-third of total pesticide mass applied to soils, which can be about 1 kg ha−1 yr−1 in the agricultural field (Barriuso et al., 2008; Boesten, 2016). Unfortunately, the incorporation of pesticides into soil organic matter does not render them benign (Bromilow, 1999). Several soil microorganisms can utilize the pesticide chemicals as sources of carbon, nitrogen, phosphorus and sulphur for their growth or the generation of energy. Sethunathan et al.

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(1969) provided the seminal report on the soil bacterium capable of degrading benzene heaxachloride, isolated from a paddy field in the Philippines. The agricultural soils have become the new hotspot for the microbial pesticide degraders as these soils receive pesticides, season after season, and year after year, especially after the agricultural intensification in the 1940s. In 1973, Sethunathan and Yoshida provided convincing evidence on the microbial degradation of organophosphorus pesticides (diazinon and parathion) in the paddy soils by Flavobacterim sp. ATCC 27551 which has been subsequently identified as Sphingobium fuliginis (Kawahara et al., 2010). Further molecular studies showed that bacterial phosphotriesterases including organophosphorus hydrolase (OPH) that help the bacteria to degrade the organophosphorus pesticides have probably evolved only during the past 70 years (Yang et al., 2003; Singh, 2009). The ‘pesticide microbiology’ with the studies on the microbial pesticide degradation has gained the global attention in the last four decades (Hill and Wright, 1978). Most pesticides, by design, act against insects and plants while the antimicrobial pesticides prevent contamination or deterioration by microorganisms such as bacteria, fungi, algae, protozoa and viruses. The antimicrobial agents encompass all the chemical compounds that inhibit or kill microorganisms (Franklin et al., 2016). The bacterial phenotypes as tolerance or persistence or resistance to pesticides or antimicrobial agents are not clearly defined. Tolerance is generally described as the ability of microorganisms to survive a transient exposure to antibiotics or pesticides. Tolerance is achieved either by the slow growth or lag (phase of bacterial growth cycle). Recently, Brauner et al. (2016) proposed the minimum duration for killing (MDK) as the quantitative indicator for tolerance to antimicrobial substances. The ‘bacterial persistence’ is observed with the emergence of a subpopulation not killed by the antimicrobial agents and characterised by the bimodal (or multimodal) time-kill curve (Gefen and Balaban, 2009). While the bacterial phenotype of tolerance or persistence to pesticides is poorly characterised, the resistance to pesticides (or antibiotics) is generally measured by the quantitative indicator of ‘minimum inhibitory concentration.’ The bacterial phenotype of pesticide degradation has nevertheless gained the attention of researchers because of the need for minimizing the adverse environmental consequences. We surmise that the bacterial pesticide-degraders are likely to have higher potential of antimicrobial resistance (or tolerance) because of the common fitness traits that require new genetical changes and molecular mechanisms. Earlier in 2003, Shafiani and Malik (2003) examined the tolerance of pesticides (endosulfan, carbofuran and malathion) and resistance to antibiotics (nalidixic acid, cloxacillin, chloramphenicol, tetracycline, amoxicillin, methicillin and doxycycline) in isolates of Pesudomonas sp., Azotobacter and Rhizobium spp. from

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agricultural fields of India. Dantas et al. (2008) provided the first report on the use of antibiotics of natural and synthetic origins as a sole carbon source by soil bacterial isolates. The recent evidence also suggests that the exposure to the commercial herbicides such as Dicamba (3,6 dichloro-2-methoxybenzoic acid), 2,4-D (2,4 dichlorophenoxyacetic acid) and Glyphosate (N-(phosphonomethyl) glycine) induced the cross resistance, i.e., adaptive multiple antibiotic resistance phenotypes in Escherichia coli and Salmonella enterica serovar Typhimurium (Kurenbach et al., 2015). However, the antimicrobial resistance as an important microbial functional trait of the pesticide-degrading microorganisms has yet to be understood, and whether induction of resistance leads to heritability is to be examined. Here, we review the microbial degradation of pesticides by addressing the types and their application in the soil and foliar environments which can become the ‘hot spots’ for the pesticide-degraders with antimicrobial resistance with the associated molecular mechanisms and genetical changes.

2. Pesticides in agriculture and public health: the regulatory stringency The registration approval of a pesticide requires critical evaluation including the environmental and toxicology studies at the national or regional level. Pesticide Properties DataBase (PPDB) of the University of Hertfordshire has the physico-chemical, toxicological, ecotoxicological and other related data for around 1150 pesticides, 700 metabolites and 100 other related substances including the country registration information and ecotoxicological data for several taxa (Lewis et al., 2016). However, the emphasis of the life-cycle concept for the pesticide management does not include antimicrobial resistance of pesticide degraders. The International Code of Conduct on Pesticide Management (also referred to as ‘Code of Conduct’) recommends the globally accepted standards, from the legislation and regulatory control to all the practices during the manufacture, transport, use and disposal. The WHO resolution WHA 63.26 urges the member states to regulate the sound life-cycle management of pesticides (WHO, 2010). The pesticide registration of US EPA has several data requirements for registration which include the scientific information on the chemistry of products and their residues, their performance, and the nature of hazard to humans, domestic animals, and non-target organisms, and their environmental fate. The registration process for antimicrobial pesticide requires additional information under Part 158W to evaluate the potential to cause unreasonable adverse effects on humans, wild life or plants, and surface water or ground water (US EPA, 2017). Despite the regulatory stringencies, the poor pesticide management is evident from the global survey, especially from the disease-endemic countries and less developed countries with lower gross domestic product. The regulation of hazardous pesticides according to the Rotterdam Convention on Prior Informed Consent (PIC) procedure is poorer in the lower income countries than in the higher income countries (Schreinemachers and Tipraqsa, 2012). There are several countries where the scientific management of pesticides is either non-existent or far below the standards currently set by the international agencies. Even in the higher income countries, the potentials of pesticides to induce antibiotic resistance have yet to be considered during the registration procedures. For example, methyl salicylate was registered as an insect repellent to be used in stored food commodities in 1996 though there were scientific reports on the induction of antibiotic resistance in the multiple-antibiotic-resistant (Mar) mutants of Escherichia coli by sodium salicylate, acetyl salicylate and salicyl alcohol (Rosner, 1985; Cohen et al., 1993; US EPA, 2005). The evolutionary imperative on the mechanism of pesticide resistance in weeds and plants due to the ‘pesticide treadmill,’ was predicted by Carson (1962). But no such prediction was made earlier on the adaptive multiple antimicrobial resistance phenotypes of microbial pesticidedegraders. Both the increasing amounts and types of pesticides applied

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in different matrices can accelerate this adaptive evolutionary trait in the microbial pesticide-degraders.

2.1. Extent of soil and foliar applications of pesticides Pre-emergent herbicides, and insecticides and fungicides to control soil-borne insects and pathogens are generally applied on the soils. Soils are also the major receptacle for the pesticides from the seed coating and foliar application, and those from crop residues, leaf fall, and root exudates. Herbicides, both in the application rates and total amounts, surpass other classes of pesticides, and likely to continue due to the decline in efficacy and the development of weed resistance and in the future climate change scenarios (Delcour et al., 2015). Benbrook (2016) reported that glyphosate (N-(phosphonomethyl) glycine) as a commercial formulation (Roundup®) was applied at about 1 kg on every hectare of the cultivated land in the USA (about 0.53 kg on all cropland worldwide) in 2014. Globally, the agricultural application of glyphosate increased from 51.3 million kg in 1995 to 747 million kg in 2014 since the approval of genetically engineered herbicide tolerant (“Roundup Ready”) soybean, maize and cotton in 1996, alfalfa in 2011 and sugar beets in 2012 (Benbrook, 2016). Atrazine, (2 chloro-4-(ethylamine)-6-(isopropyl amine)-s-triazine), another commonly used herbicide, has an annual agricultural consumption of N3 million kg in the USA, though banned in the European Union since 2003 (Pathak and Dikshit, 2012). In 2014, the US EPA (2014) approved the combined use of 2,4-D and Roundup on the genetically engineered Enlist soybean and Enlist corn. Likewise, the approval of high-yielding genetically engineered ‘Roundup Ready 2 Xtend’ soybeans and ‘Bollgard II XtendFlex’ cotton has led to increased application of dicamba along with glyphosate. The combined use of different herbicides might exacerbate the soil contamination with increasing acreage under the genetically engineered, herbicide tolerant cultivars. Hydrophobic organochlorine pesticides generally have log Kow values between 5.5 and 6.9. Their application has led to the soil contamination with low to moderate levels and with the breakdown products for several years later (Thomas et al., 2008). The initial persistent organic pollutants (POPs), recognized by the United Nations Environment Programme, have nine pesticides (aldrin, chlordane, DDT, dieldrin, endrin, heptachlor, hexachlorobenzene, mirex and toxaphene) of a total 12 chemicals. The additional 16 new POPs include seven pesticides (α-hexachlorocyclohexane (α-HCH), β-HCH, lindane (γ-HCH), chlordecone, pentachlorobenzene, pentachlorophenol and its salts and esters, and technical endosulfan and its related isomers), being five of them organochlorine insecticides (UNEP, 2017). Persistence of chemical compounds will increase the probabilities of developing capabilities, not only to degrade pesticides in microorganisms but also to develop cross-resistance against antimicrobial agents. Several systemic pesticides are registered for use in soil and they can enter plants by roots, stem or leaves, and are translocated to other parts. In addition, the coformulation with polymers or the use of nanomaterials can increase the systemic action of these pesticides. Neuro-active, nicotine-based systemic insecticides (neonics) such as neonicotinoid and fipronil were introduced commercially in the mid1990s. They are extensively used globally, toxic at very low doses, and have higher persistence in soil and water (Furlan et al., 2018). The multiple routes of exposure to neonics can result in chronic effects in different species while the acute effects can be 5000 to 10,000 times more toxicity than DDT (Tsvetkov et al., 2017). The ‘magic nitro’ (nitroimine, _N\\NO2) group contributes to high levels of selective toxicity of neonicotinoid insecticides (e.g., imidacloprid and thiamethoxam) for insects over vertebrates, but its slight alterations (guanidine/desnitro derivates) can reverse its selective toxicity (Tomizawa and Casida, 2005). The microbial degraders include pure bacterial cultures or consortia capable of using neonicotinoids as a sole source of carbon or nitrogen for growth or by cometabolism (Hussain et al., 2016).

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The global foliar surface is estimated to be N108 km2 and the foliar microbiome is estimated to have about 1016–1018 bacteria on per km basis (Penuelas and Terradas, 2014). Although only a fraction of the global foliar surface areas belongs to agricultural crops, they are cultivated season after season, and with several sprayings of pesticides. Despite the efforts to improve the foliar uptake (Wang and Liu, 2007), contact or systemic pesticides will impact the foliar microbiome significantly. Several bacteria are reported to produce toxins naturally. The presence of pesticides can increase the chances of developing adaptive multiple antimicrobial resistance, especially in those which can degrade the pesticides. Because of intensive application, accentuated by their chemistries, persistence, and tendencies for the long-range transboundary atmospheric transport, pesticides can be detected around the world, even in places where they are not applied to control the pests. Both the soil and foliar matrices have significantly higher chances of pesticide-degraders with antimicrobial resistance as an additional functional trait. 2.2. Obsolete pesticides Though the agricultural soils are the major recipient of pesticides applied, the storage and use of registered and obsolete (including deregistered locally or banned internationally) pesticides, deteriorated products contribute to the soil contamination. The FAO's programme on the Prevention and Disposal of Obsolete Pesticides suggests a stockpile of 3.3 × 106 kg in India (2004) while the largest stockpiles of 241 × 106 kg in the eastern European countries in 2006 (FAO, 2017). The obsolete pesticides are estimated to about 2170–111,200 t in some OECD and partner countries (OECD, 2014). Most of the stockpiles are poorly stored and the contents leak to contaminate the soils. The collection and disposal costs are around 3500–5000 US dollars, depending on the size of project and location (CropLife International, 2017). Ten years after the launch of Africa Stockpiles Programme (ASP), only about 4000 t have been treated as against 50,000 t of obsolete pesticides which are estimated to be in Africa. The obsolete POPs such as DDT, Toxaphene, Mirex and dieldrin have half-lives of 22–30 years, about 14 years, about 12 years, and up to 7 years, respectively (Pesticide Action Network Europe, 2010). Since the expertise and resources to collect, safeguard and dispose the obsolete pesticides are hard to find in many countries, the storage of these pesticides along with the application of registered pesticides will be one of the major drivers of soil pollution, especially with adverse effects on the soil microbiomes. Even the changes in agricultural practices can influence persistence and transfer dynamics of residual pesticides in soils. Sabatier et al. (2014) reported that application of postemergence herbicide, glyphosate, induced soil erosion, leading to the release of previously persistent DDT residues from vineyard soil in France. Longer the persistence of obsolete pesticides in soils, greater are the chances of developing capabilities to degrade them by microorganisms. 3. Microbial degradation of pesticides: a means of cleansing the soil environment Pesticides undergo several interdependent chemodynamic processes which determine their persistence, distribution and degradation in the environment (Table 1). The pesticide transformation includes both chemical and biological degradation. The chemical degradation generally involves oxidative, reductive, hydrolytic or photolytic reactions while the biological degradation is due to oxidation, reduction, hydrolysis and conjugation (Gan et al., 2002; Zacharia, 2011; Megharaj et al., 2011; Hou et al., 2018). The IUPAC (1993) states biodegradation as “breakdown of a substance catalysed by enzymes in-vitro or invivo.” For the hazard assessment of chemicals, biodegradation is described as primary, environmentally acceptable, and ultimate depending on the alteration, removal of undesirable properties and complete breakdown, respectively. ECHA (2017) defines biodegradation as “the

Table 1 Chemical properties and parameters of pesticides accentuating environmental risks. Property

Parameter(s) measured

Solubility in water (g L−1) Volatility (Pa) Henry's constant Binding soil/sediments (g kg−1) Partitioning Adsorption, deadsorption (KD) Octanol-water Kow, Koc Binding KD to media Mobility

Persistence

Dissipation

Environmental risk(s) Leaching Run-off potential Volatility Air pollution Bioaccumulation Release from run-off sediments Inability for containment Accumulation/transport Potential for accumulation

- Half-life in soil/water/foliage Degradation

-

Chemical hydrolysis (per sec) UV degradation (per sec) pH effect Rate of biodegradation (per sec)

Adapted from Kennedy et al. (1998).

biologically-mediated degradation or transformation of chemicals usually carried out by microorganisms.” The microorganisms can detoxify the pesticides by converting the chemicals into nontoxic metabolites, mineralize by complete degradation, and cometabolize them while growing on other compounds, or even activate the non-toxic chemicals to toxic chemicals (Singh et al., 1999). Audus (1949) provided the first report on biodegradation of 2,4-D in garden loam soil, attributing to the microbial activities almost entirely. Several strains belonging to different microbial genera capable of degrading this herbicide have been isolated, with the catabolic pathways of involving either dioxgenase or dehalogenase before the acetoxyl group (Huang et al., 2017). But, dichlorphenol, which can form due to its incomplete degradation, is more toxic than 2,4-D. Parathion (O,O-diethyl-O-p-nitrophenyl phosphorothioate), an organophosphate (OP) pesticide, was found to undergo rapid degradation in rice soils either via reduction of the nitro group or via hydrolysis at the PO-C linkage after repeated applications (Sethunathan, 1973; Sethunathan and Yoshida, 1973; Siddaramappa et al., 1973; Reddy and Sethunathan, 1983). In the diazinon-retreated rice fields, the acclimatized microorganisms proliferated and rapidly degraded diazinon; a Flavobacterium sp. ATCC 27551 (S. fuliginis) was isolated, which could hydrolyse diazinon and mineralize its pyrimidinyl moiety to CO2 (Sethunathan and Yoshida, 1973). The significance of enhanced biodegradation was realised world-wide since late 1970s as the accelerated biodegradation of carbamate pesticides led to failures of rice brown planthopper control (Venkateswarlu and Sethunathan, 1978, 1979) and rootworm control in the corn fields of the USA and Canada (Fox, 1983). HCH, an organochlorine insecticide containing a mixture of α-, β-, γ-, and other isomers was found to undergo rapid degradation under anaerobic conditions but not under aerobic conditions for long (MacRae et al., 1967). The aerobic biomineralization of α-HCH isomer was reported from the HCH-contaminated site in the Netherlands by Bachmann et al. (1988) and of α-, γ- and stable β-isomers by a Pseudomonas sp. from the sugarcane rhizosphere soil from India by Sahu et al. (1990). Contrary to the common belief on the recalcitrance of HCH isomers to aerobic degradation, these isomers are susceptible to aerobic biodegradation after two or three application in pure cultures of bacteria, in soils, and also in the HCH-acclimatized soils (Sahu et al., 1990). Bhuyan et al. (1993) reported not only enhanced biodegradation but also a distinct build-up of aerobic, but not anaerobic, microorganisms capable of degrading γ-HCH upon its repeated additions. The degradation of pesticides by microorganisms can occur metabolically using pesticides as a source of energy or nourishment, or cometabolically. The degradation is generally described using the first

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order kinetics since the microorganisms or enzymes involved are present in excess compared to the pesticide molecules. Because of heterogeneous nature of the soil medium where degradation can occur at different rates within individual regions of the soil and the available fraction of pesticide can decrease with time due to slow sorption and diffusion processes, the First Order Multi Compartment (FOMC) equation and the Double First Order in Parallel (DFOP) equation are proposed to describe pesticide degradation (Borgesen et al., 2015). The limitations on the availability as substrates and constraints on the microorganisms for sensing and uptake of pesticides or antimicrobial substances are many in the soil environments. 3.1. Chemotaxis: the foremost step for bacterial degradation of pollutants Bioavailability of pesticide chemical is one of the major limitations for microbial degradation (Head, 1998). Semple et al. (2004) defined the bioavailable compound as that which is freely available to cross an organism's cellular membrane from the medium the organism inhabits at a given time. The hydrophobic nature of most pesticides and bacterial membranes necessitate the dissolution of the chemicals in the aqueous phase or the movement of bacteria to the nonaqueous-phase liquid (NAPL)-water interfaces for the direct adhesion to the hydrophobic chemicals. If the dissolution rate of a chemical is more than the growth rates of bacteria, the potential for biodegradation of that chemical is generally higher. Chemotaxis is an important mechanism which allows bacteria to move toward (chemoattraction) or away (chemorepulsion) from chemicals (Pandey and Jain, 2002; Lacal et al., 2012). There are two categories of chemotaxis: metabolism-independent and metabolism dependent chemotaxis. In the metabolism-independent chemotaxis, both the metabolizable and nonmetabolizable analogues are the attractants and transmembrane chemoreceptors (chemotaxis transducers) determine the movement of bacterial cells (Bren and Eisenbach, 2000). On the contrary, the metabolizable chemicals are the only attractants in the metabolism-dependent chemotaxis (Alexandre et al., 2000). Parales et al. (2000) reported that the toluene-degrading Pesudomonas putida F1 when grown on toluene induced chemotaxis toward toluene, benzene, and trichloroethylene. Even the P. putida F1 mutants of todST genes adjacent to the tod structural genes (todRXFC1C2BADEGIH) that were defective in toluene dioxygenase and did not grow in toluene were chemotactic toward toluene, suggesting the involvement of metabolism-independent chemotaxis mechanism. The chemotaxis was induced in Ralstonia eutropha JMP123(pJP4) when grown on 2,4D (Hawkins and Harwood, 2002). But, the tfdK mutant with the catabolic plasmid pJP4 having tfd cluster but defective nonessential permease TfdK for growth on 2,4-D was nonchemotactic. Samanta et al. (2000) isolated the strain SJ98 of Ralstonia sp. which was chemotactic toward the metabolizable nitroaromatic compounds such as p-nitrophenol and o-nitrobenzoate but not the nonmetabolizable chemicals such as phenanthrene and p-nitroaniline. Generally, mutation in the metabolic degradation pathways of pesticides does not affect the chemotaxis of bacteria possessing the metabolism-independent mechanism while it does in bacteria with metabolism-dependent mechanism. The diversity of signal transduction mechanisms including hyperchemotaxis or chemotactic repellence can determine persistence or the kinetics of pesticide degradation (Krell et al., 2011; Lacal et al., 2012). 3.2. Membrane transport systems: the basis for xenobiotics uptake Transduction of chemotactic signals involves chemoreceptors which are cytoplasmic membrane-bound proteins; the chemoreceptors for new, structurally similar xenobiotic chemicals evolve from the existing chemoreceptors (Parales et al., 2015). Increased number and types of methyl-accepting chemotaxis proteins (MCPs) contribute to the metabolic diversity of soil bacteria. The expression of bacterial genes related to transport, chemoreceptor and degradation is generally coordinated

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for optimal growth and survival. The plasmid-encoded major facilitator superfamily (MFS) transporter (TfdK) for the entry of 2,4-D has been detected in R. eutropha JMP(pJP4) (Hawkins and Harwood, 2002). The transporter, Tap (an MCP), is constitutively expressed in E. coli which detects dipeptides and pyrimidines. Because of structural similarity of pyrimidines with s-triazines, the uptake of atrazine and related s-triazines was reported in E. coli (Liu and Parales, 2009). Based on the size of the ligand binding region (LBR) of MCPs, there are two clusters: Cluster I MCPs and Cluster II MCPs (Lacal et al., 2010). The diversity of chemoreceptors and the acquisition of their genes by transmissible plasmids may increase the signal sensing abilities in Bacteria and Archaea (Wuichet and Zhulin, 2010). The presence of outer membrane offers additional protection in terms of resistance to antimicrobial agents or pesticides in Gram-negative bacteria than in Gram-positive bacteria (Epand et al., 2016). Both the entry of any chemical and the excretion of metabolic end products from the cell depend on the cytoplasmic membrane transport systems. Transport proteins (porters), which are even stereospecific carriers mediate the trans-membrane solute translocation involving solute-non-specific, and solute-specific channels with different energy coupling mechanisms. Paulsen et al. (1998a, 1998b) and Lorca et al. (2007) observed that the eubacterial genome size is an important determinant and N9% of total encoded genes correspond to the transport systems. The genes for transporters constituted 21.7% in the genome of Pseudomonas silesiensis sp. nov. strain A3T isolated from the pesticide sewage treatment plant at Jaworzno City, Poland (Kaminski et al., 2017). The transporters of the major facilitator (MF) superfamily (of secondary active transporters) and those of ATP-binding cassette (ABC) superfamily (of primary active transporters) are diverse, largest, and capable of transporting low molecular weight compounds, and of low as well as macromolecules, respectively. The presence of large number of paralogs in bacterial genomes contributes to the broad specificities and variable polarities for the transporters of the ABC and MF superfamilies. In a recent report, Fischer et al. (2010) showed that the microbial uptake of low-molecular-weight organic substances (14Clabelled glucose, alanine and acetate as test chemicals) out-competed the sorption (physicochemical process) of alanine and glucose in soil. Transport capabilities, in terms of distribution and types, are critical to the metabolic capabilities of microorganisms including pesticide degradation or xenobiotic resistance. Water-soluble pesticides (e.g., methomyl, diuron and imidacloprid) have generally low persistence in the soil environment. The formulation, mode and methods of their application necessitate new strategies. Sun et al. (2014) encapsulated methomyl using amphiphilic biocopolymers in shell cross-linked nanocapsules for its controlled release. The bacterial membranes have porins that form channels to facilitate the uptake of hydrophilic compounds of certain size limits, can be up to 106 copies per cell and have diverse roles including that of receptors for antibodies, bacteriocins and bacteriophages (Achouak et al., 2001; Fernandez and Hancock, 2012). Salicylate, when tested at 1, 5 and 8 mM, regulated the expression of the genes for OmpC and OmpF porin proteins differentially in Serratia marcescens (Begic and Worobec, 2006). There are several salicylate containing pesticides commercially available (e.g., monotributyltin salicylate). Interestingly, salicylate is also reported to promote the expression of marRAB operon to induce multiple-antibiotic-resistance in E. coli (Alekshun and Levy, 1999). Kurenbach et al. (2015) tested the hypothesis whether the reduced synthesis of outer membrane porins contributed to the antibiotic susceptibility of E. coli and Salmonella enterica serovar Typhimurium when exposed to the sublethal concentrations of Dicamba, 2,4-D and Glyphosate. For the selected strains, the lethal concentrations of these pesticide formulations were N13,000, 4500 and 6100 ppm acid equivalents of Kamba500 which contained 500 g dimethyl salt of dicamba L−1, 2,4-D amine 800 WSG with 800 g dimethylamine salt of 2,4-D, and Roundup that contained 360 g isopropylamine salt of glyphosate L−1, respectively. Thus, the toxic concentrations of pesticides to

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microorganisms are less, relative to those of the natural or synthetic antibiotics. When the pesticides and antibiotics were provided together, the antibiotic susceptibility of bacterial test strains varied characteristically depending on herbicide, and antibiotic, in a dose-dependent manner, and the minimum inhibitory concentrations of antibiotics increased or decreased. Though Kurenbach et al. (2015) tested the herbicides at the sublethal concentrations (≤1950 ppm Kamba and 2,4-D, and ≤1240 ppm Roundup), the applications of these pesticide formulations at the fields would result in the concentration levels of b25 ppm. However, the surfactants and other organic and inorganic chemical compounds as contaminants in the commercial formulations of pesticides might accentuate the toxic effects on the microorganisms. Innate capabilities for uptake or resistance to hydrophilic xenobiotics including pesticides and antibiotics may depend on the numbers and types of porins in the cell membranes. Increased or reduced synthesis of porins will alter the uptake of nutrients and xenobiotics (pesticides, antibiotics, or antimicrobial substances) in microorganisms and subsequent degradation potential for pesticides or resistance to antibiotics/antimicrobial substances. The porin proteins have now been employed to develop the biotechnological tools such as biosensors. Yan et al. (2013) developed a novel acetylcholinesterase (AChE) biosensor (AChE liposomes bioreactor) by embedding porins in the lipid membrane of multilayer films containing multiwall carbon nanotubes and chitosan to detect an organophosphate pesticide, dichlorvos at a concentration ranging from 0.25 to 1.75 μM and from 2.00 to 10.00 μM with a detection limit of 0.68 ± 0.076 μg L−1. 3.3. Efflux pumps: role in antimicrobial resistance Microorganisms possess efflux pumps to pump out the toxic substrates which have gained entry into the cytoplasm as well as those toxic, endogenous metabolites out of the cytoplasm (Levy, 1992). Nearly 6 to 18% of all transporters present in a bacterium belong to efflux pumps. The bacterial efflux pumps are classified into five major families: (i) resistance-nodulation-division (RND) family, (ii) MF superfamily, (iii) ABC superfamily, (iv) small multidrug resistance (SMR) family and (v) multidrug and toxic compound extrusion (MATE) family (Blanco et al., 2016). Both Gram-positive and Gram-negative bacteria have all these transporters except those of RND family, which are present only in Gram-negative bacteria. The genes encoding proteins belonging to specific efflux pumps (for a specific substrate) are generally on the plasmids while those of multi-resistance systems (for several, structurally dissimilar compounds) are on the chromosomes (Paulsen et al., 1998a, 1998b; Fernandez and Hancock, 2012). Most efflux pumps can transport multiple substrates, are chromosomally encoded and tightly regulated, and confer multidrug resistance in bacteria. In an earlier report, the overexpression of antibiotic efflux pumps (MexAB-OprM efflux system) was attributed to the organic solvent tolerance of Pseudomonas aeruginosa and the development of antibioticresistant phenotype of P. aeruginosa in the presence of hexane (Li and Poole, 1999). Overexpression of efflux pumps due to the presence of biocides is known to select for antibiotic-resistant mutants (Fraise, 2002). Kurenbach et al. (2015) reported the increased tolerance of E. coli to chloramphenicol and kanamycin in the presence of dicamba and glyphosate, respectively, by employing the growth assays with an efflux pump inhibitor [Phe-Arg β-naphtylamide (PaβN)] for the AcrAB efflux pump. Efflux pumps may play a significant role alone or in association with other membrane proteins or porins in developing increased antibiotic or antimicrobial resistance, in the presence of pesticides. 3.4. Plasmid and chromosome encoded catabolic genes: role in pesticide biodegradation

microorganisms (Weinstock, 2001). Chakrabarty (1972) showed the genetic basis of salicylate degradation involving the SAL plasmid in Pseudomonas putida R1. Five years later, Pemberton and Fisher (1977) gave the first report on the plasmid encoding genes for degradation of 2,4-D. A 58-megadalton conjugal plasmid, pJP1, was found to encode the degradation of 2,4-D and 2 methyl-4-chlorophenoxyacetic acid but not phenoxyacetic acid in Alcaligenes paradoxus (Fisher et al., 1978). The plasmids (pJP3, pJP4, pJP5, and pJP7) of Alcaligenes eutrophus encoded the degradation of m-chlorobenzoate and conferred resistance to merbromin, phenyl mercury acetate and mercuric ions, and can transfer freely to E. coli, Rhodopseudomonas sphaeroides, Rhizobium sp., Agrobacterium tumefaciens, Pseudomonas putida, Pseudomonas fluorescens and Acinetobacter calcoaceticus (Don and Pemberton, 1981). The catabolic genes (tfdA, tfdB, tfdC, tfdD, tfdE, and tfdF) for 2,4D degradation were identified in the plasmid pJP4, which also encoded mercuric resistance, of Alcaligenes eutrophus JMP134 (Don and Pemberton, 1985). The plasmid mediated conjugation process can transfer the catabolic genes for 2,4-D degradation to other soil bacteria in nature (DiGiovanni et al., 1996; Newby et al., 2000a). Sakai et al. (2014) reported the distribution of pM7012 (a megaplasmid group) which encodes 2,4-D degradation and arsenic resistance in strains of Burkholderia and Cupriavidus isolated from Japan and the United States. For the organophosphate pesticides, the involvement of plasmid (pCS1) was first reported in Pesudomonas diminuta for parathion hydrolysis (Serdar et al., 1982). The opd gene encoding the organophosphorushydrolysing (OPH) enzyme was detected on different plasmids of Flavobacterium sp. strain ATCC 27551 and Brevundimonas diminuta MG (Harper et al., 1988; Mulbry and Karns, 1989). Catabolic genes are also found to be chromosomally encoded (e.g., tfdARASC which encodes for 2,4-D/α-ketoglutarate dioxygenase in degradation of 2,4-D in Burkholderia sp.) (Suwa et al., 1996). The horizontal transfer of chromosomal tfdA genes in different species of Burkholderia was demonstrated by Matheson et al. (1996). Horne et al. (2002) reported the chromosomally encoded opd gene in Agrobacterium radiobacter P230. Several catabolic genes have been identified from bacterial and fungal organisms (Ortiz-Hernandez et al., 2013). In the microbial metabolism of atrazine metabolism in soils, the gene copies of atzD (cyanuric acid hydrolase) have been found to be correlated with its degradation in the drilosphere (soil bioturbated by earthworms) of agricultural fields (Monard et al., 2010). Sherchan et al. (2013) employed the quantitative polymerase chain reaction for the gene copies of atzA, which codes for the key enzyme atrazine chlorohydrolase that dechlorinates atrazine to dealkaylated metabolites for detection of atrazinedegrading bacteria as an indicator of pesticide contamination in the coastal waters. Genome sequencing provides new insights into the evolution of these catabolic genes (Copley et al., 2012; Martini et al., 2016; Nielsen et al., 2017). High-throughput-sequencing of the plasmid pool (mobile genetic elements) from the biopurification systems, which were constructed for the elimination of pesticides from wastewaters of farm activities, led to the identification of genes encoding enzymes involved in pesticide and hydrocarbon degradation as well as genes conferring resistance to antimicrobial substances (Martini et al., 2016). The gene transfer between chromosomes and plasmids and horizontal gene transfer (HGT) for the gene clusters sdpA and cadA and the adaptation for chlorophenoxy pesticides were inferred from the complete genome analysis of Sphingobium herbiciovorans (Nielsen et al., 2017). New genetic and genomic tools will advance our understanding on the evolution and adaptation of microorganisms with the genes encoding enzymes for degrading synthetic pesticides and other novel functions including antibiotic/antimicrobial resistance. 3.5. Enzymes for microbe-mediated pesticide degradation

Heritable information by genes and their regulation and expression has gained the global attention because they help to decipher the functional and metabolic pathway organization of cellular processes in

Pesticides, if they are not pumped out by the efflux pumps, are enzymatically transformed in the cytoplasm. The microorganisms are also

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capable of transforming pesticides by extracellular enzymes. Several microbial enzymes such as hydrolases, oxygenases, dehydrogenases, esterases, lyases, phosphatases, and dehalogenases are reported to be involved in pesticide degradation (Scott et al., 2008). Most hydrolases are extracellular and cleave amide, peptide bonds, carbon-halide bonds, and ester bonds in anilides, phenyl-ureas, esters of carbamic, thiocarbamic, phosphoric, and thiophosphoric acids. The parathion hydrolase producing Flavobacterium sp. was isolated from the paddy fields in the Philippines (Sethunathan and Yoshida, 1973). Oxygenases are of two groups, monooxygenase requiring reduced pyridine nucleotides as cofactors and dioxygenases which do not require any reduced compound as a cofactor. Hydrolases and oxygenases are the two major groups of microbial enzymes investigated for pesticide degradation. The biodegradation reactions have been formalized and catagorized which have led to the development of “Metarouter,” and the University of Minnesota Biocatalysis/Biodegradation Database (UM-BBD) (Pazos et al., 2005; Gao et al., 2010). Cuesta et al. (2015) suggested that the microbial adaptation for pesticide resistance or degradation can drive innovation, exchange, and demise of enzyme function. Innovation in enzyme function can involve point mutations, creating a new activity in an existing enzyme. The innovation-amplification-divergence model explains this type of evolution of neo- and subfunctionalization of enzymes with adaptive cycles (Copley, 2012). Besides, the combination of chemistry-driven and substrate-driven evolution can also make enzymes to accommodate alternative chemistries such as transferases becoming oxidoredutases, hydrolases and lyases. Many enzymes are promiscuous, capable of performing multiple reactions, either due to divergent evolution (i.e., functional diversification of enzyme superfamilies) or convergent evolution (similar active sites in analogous enzymes) (Galperin and Koonin, 2012; Cuesta et al., 2015; Table 2). The transfer of genes encoding enzymes between organisms (horizontal gene transfer or endosymbiotic transfer), gene duplication or loss, and gene fusion are important evolutionary mechanisms of biosynthetic diversification (Michael, 2017). A better understanding of proteomics and functional genomics is critical to gain insights on the global metabolic and regulatory networks, especially in the key microbial players responding to the pesticide applied in the environments. Besides the enzymes within microbial cells, free enzymes or immobilized extracellular enzymes are present in soils. Pesticides may have negative effect on soil enzymes such as hydrolases, oxidoreductases and dehydrogenases (Menon et al., 2005; Tu, 1992) or positive effect (Megharaj et al., 1999). The genetic analysis of catabolic genes encoding enzymes associated with pesticide degradation is of importance in order to assess the distribution of this capability amongspecies as well as across the environments. 3.6. Evolution and mechanisms of inheritance for microbial degradation The evolution of highly efficient enzymes in bacterial systems was attributed to: (i) population size, (ii) mutation rate, (iii) short generation times, (iv) horizontal gene transfer, and (v) large biochemical repertoire of diverse organisms (Russell et al., 2011). The selection of an existing genetic trait, among the highly variable traits of the gene pool, is an important mechanism for genes that confer resistance or those encoding catabolic enzymes (Bergman, 2003). The repertoire of signal transduction systems, membrane transport systems (transporters, porins, and efflux pumps), and catabolic genes make the microorganisms intrinsically capable of utilizing or resisting the chemicals. Spontaneous occurrence of mutations can improve the phenotypic variability, enabling to adapt to the pressure of pesticide chemicals, in the bacterial systems. This mutational selection can be transmitted, and the new traits become stable in the presence of ‘evolutionary pressure.’ Besides, the new traits (genes in plasmids, integrons, transposons, naked DNA and others) can be acquired through conjugation, transduction and transformation of the horizontal gene transfer mechanisms (DiGiovanni et al., 1996; Newby et al., 2000b). In addition to the

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Table 2 Enzymes involved in pesticide degradation and antibiotic inactivation. Enzyme class

Pesticide degradation (Reference)a

Oxidoreductases Oxidoreductase involved in endosulfan degradation Monooxygenases Dioxygenases Mixed function oxidases (Sutherland et al., 2002; Scott et al., 2008) Transferases Glutathione S-Transferase (Vuilleumier, 1997)

Hydrolases

Lyases

a

Organophosphorus hydrolase Phosphotriesterases Acetylchloinesterases Haloalkane dehalogenases (Grimsley et al., 1997; Bigley and Raushel, 2013; Singh and Walker, 2006) Nonheme iron-dependent Dioxygenase with C as lyase activity (Bugg and Ramaswamy, 2008)

Antibiotic inactivation (Reference)a TetX (oxidation of tetracycline) Monooxygenase inactivation of rifamycin (Markley and Wencewicz, 2018; Liu et al., 2018)

Aminoglycoside acetyltransferases Chloramphenicol acetyltransferases Streptogramin acetyltransferases Phosphotransferases Aminoglycoside kinases Thioltransferases Nucleotidyltransferases ADP-ribosyltransferases Glycosyltransferases (Shaw and Leslie, 1991; Thal and Zervos, 1999) β-lactamases, macrolide esterases and epoxidases-mediated inactivation of β-lactams and macrolides (Bush and Jacoby, 2010; Morar et al., 2012) Vgb for Type B streptogramin resistance (Mukhtar et al., 2001)

Selected references are included.

intrinsic and acquired mechanisms, an adaptive mechanism also exists, which is transient in nature and the removal of the inducing condition can revert the alterations in gene and/or protein expression, as in the case of antibiotic resistance (Fernandez and Hancock, 2012). The catabolic functions are also reported to arise independent of evolution, such that genes with different sequences may mediate the same reaction as in the case of phosphotriesterase (PTE), methyl parathion hydrolase (MPH), organophosphorus acid anhydrolase (OPAA), diisopropylflurophosphatase (DFP), and paraoxonase 1 (PON1) (Bigley and Raushel, 2013). This array of enzymes differs in their protein sequences, three-dimensional structures, and catalytic mechanisms but with the common functional trait. Intrinsic, acquired or adaptive mechanisms for new traits and those functional traits independent of evolution for tolerance or resistance or utilization of various pesticide chemicals may get mediated due to the selection pressure. 4. Blind spots to hotspots: antimicrobial resistance of pesticidedegrading microorganisms The soil microbiome has a rich repository of genes to repel, tolerate, resist or degrade chemicals of natural or synthetic origins, largely due to the diverse chemicals present in the soil environments. There exists an enormous chemical diversity among the natural antibiotics in soils. The actinobacterial members of the soil microbiome produce several clinically relevant antibiotics. These actinobacteria also possess resistance genes, which are critical for self-protection and are often clustered in the antibiotic biosynthetic operons (D'Costa et al., 2006; Wright, 2007). The natural antibiotics at low doses are known to have other biotic functions which include the expression of genes regulating virulence, colonization, motility, gene transfer, and secondary metabolite production (Aminov, 2013). The biosynthetic pathways for antibiotics such as erythromycin, streptomycin and vancomycin has been

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considered to have evolved around 240–880 million years ago (Baltz, 2007). Hall and Barlow (2004) suggested that the evolution of serine β-lactamases which can cleave the β-lactam of the β-lactam antibiotics was about two billion years ago and some of these enzymes on plasmids for millions of years. These evolutionary events imply the obvious development of antibiotic resistance genes (resistome) in the producers to protect themselves during the same period. The common mechanisms of antibiotic resistance in bacteria include: (i) antibiotic inactivation by hydrolysis, group transfer and redox process; (ii) target modification such as peptidoglycan structure alteration, protein synthesis interference and DNA synthesis interference; (iii) efflux pumps and outer membrane permeability changes in order to reduce the concentration, without modification of the compound itself; or (iv) target bypass-some bacteria become refractory to specific antibiotics by bypassing the inactivation of a given enzyme (Dzidic et al., 2008). Because of the RND superfamily transporters (efflux system), the Gramnegative pathogens are intrinsically resistant to many antibiotics (Li and Nikaido, 2004). In general, the soil antibiotic resistome consisting of both antibiotic- and cryptic resistance genes from pathogens, opportunistic pathogens, and non-pathogenic bacteria is comprehensive, adaptable and extensive (Wright, 2007). In the recent times, the environmental burden of chemical pesticides and antimicrobial substances is increasing with the intensification of agricultural, industrial and societal activities. Both the classes of chemicals are present together in certain environments (Kemper, 2008). The evolution of resistance to antibiotics/antimicrobial substances including pesticides may now depend on the types and bioavailable concentrations of these substances in the natural environments. In the oligotrophic environments such as soils and foliage, microorganisms might employ the strategy of ‘multivorous way of life,’ by metabolizing different carbon substrates simultaneously (Egli, 2010). This ‘mixed substrate growth’ may include substrates such as pesticides, antibiotics and antimicrobial substances, giving the kinetic advantage and metabolic flexibility. The scientific evidence suggests the ‘multivorous way of life’ in microorganisms isolated from soils. D'Costa et al. (2006) constructed a library of 480 strains, screened them against antibiotics of natural and their semisynthetic derivatives and synthetic molecules, and found that every strain showed multi-drug resistance to seven or eight antibiotics on average. What is more interesting is the report of Dantas et al. (2008) which demonstrated the capabilities of soil bacterial isolates to grow on antibiotics of natural and synthetic origin (13–17 of eight major classes) at a concentration of 1 g L−1 as a sole carbon source. Each of the antibiotic-subsisting isolate was found to be resistant to multiple antibiotics at clinically relevant concentrations suggesting the soil microorganisms as an important source of multiple antibioticresistance machinery. Nevertheless, the bacterial catabolism of antibiotics (Penicillin, Carbenicillin, Streptomycin and Trimethoprim) was challenged by Walsh et al. (2013) who demonstrated the initial survival of bacterial isolates on high concentrations of antibiotics but no degradation of streptomycin or trimethoprim. The multidrug-resistant soil bacteria containing resistance cassettes against β-lactams, aminoglycosides, amphenicols, sulphonamides and tetracyclines were demonstrated by Forsberg et al. (2012). These resistance cassettes had perfect nucleotide identity to genes from the human pathogens. In the nature, the bacterial responses to the presence of antibiotics or antimicrobial substances might vary from tolerance, subsistence, dependence to even the persistence of biofilms (AmabileCuevas, 2013). Way back in 1985, Rosner demonstrated the induction by salicylates and other chemotactic repellents on the nonheritable resistance to chloramphenicol and other antibiotics in E. coli K-12. The exposure to the commercial herbicides such as dicamba, 2,4-D and glyphosate was found to induce the cross resistance, i.e., adaptive multiple antibiotic resistance phenotypes in Escherichia coli and Salmonella enterica serovar Typhimurium (Kurenbach et al., 2015). Rivera-Ramirez et al. (2016) isolated 333 pesticide resistant mutants of Salmonella enterica serovar Typhimurium, of which 236 mutants

showed reduced susceptibility to amphicillin, chloramphenicol, gentamycin, trimethoprim-sulfamethoxazole and nitrofurantoin. Several of these non-inherited resistant phenotypes in microorganisms are difficult to detect with the routine methods. Interestingly, even the agricultural application of nitrogen was found to influence the content of antibiotic resistance genes (ARGs) in soils (Forsberg et al., 2014). It is not known whether copiotrophs evolve faster than oligotrophs in developing ARGs. The agricultural practices such as the application of manure or dairy wastewater can enhance the spread of antibiotic resistance genes in the farms (Udikovic-Kolic et al., 2014; Ruuskanen et al., 2016; Dungan et al., 2018; Kuppusamy et al., 2018). Nordenholt et al. (2016) demonstrated that the application of manure collected from a barn containing deep straw bedding for swine increased the half-life of atrazine (500 μg kg−1) applied to a sandy loam soil by N10 days but decreased its degradation by 22% and its mineralization by 50%. The combined presence of pesticides and antibiotics in manures and wastewaters will alter the responses and the composition of microbial communities, which will require high resolution methods such as metagenomic analysis for monitoring. Lau et al. (2017) employed the functional metagenomics to study the resistome of Canadian agricultural soils and identified 34 new ARGs, several of them encoding for (multi)drug efflux systems. The multidrug resistance (MDR) efflux pumps are also reported to be involved in detoxification of metabolic intermediates, virulence and signal trafficking including the extrusion of substrates such as antibiotics, heavy metals, organic pollutants, quorum sensing signals and many others (Blanco et al., 2016). There are suggestions that the resistance to single classes of antibiotics has developed to multidrug resistance (MDR) and extensive drug resistance (XDR) since the 1990s (CDC, 2006). Increased mutation rates which can lead to an accelerated rate of selection or development of cross resistance (mutation in the presence of a chemical or drug leads to the resistance or tolerance of another chemical or pesticide or drug) might increase the resistance determinants in the soil antibiotic/antimicrobial resistome (Fig. 1). Since the evolutionary processes are influenced by several unknown factors, the prediction of pesticidedegrading microorganisms with antimicrobial resistance requires the rules of microbial fitness in different environments with diverse chemicals be defined. In addition, the current definition of ‘minimum inhibitory concentration’ (MIC) for antimicrobials is based on the inhibition of visible growth of a microorganism (Andrews, 2001). Target species, and chemicals and their concentrations are important determinants of MICs. Several pesticides have mixed or no direct effects on bacteria, while insecticides, herbicides and fungicides adversely affect algal and fungal species (Staley et al., 2015). But, the toxicity of pesticides to microorganisms is difficult to determine because high or low concentrations may switch on and off different genes, and have either harmful or beneficial effect on the microbial functions. Several pesticides have been tested against pure cultures of microorganisms at high concentrations (Table 3). With high concentrations of several pesticides having ‘no effects,’ the induction of cross resistance for antimicrobial agents is very likely in the pesticide-degrading microorganisms. Altogether, the recent findings strongly support the premise that the pesticide degrading microorganisms will have enhanced capabilities of antimicrobial/antibiotic resistance (Kurenbach et al., 2015; Lau et al., 2017). 5. ‘One Health’: human implications of antimicrobial resistance in pesticide degraders Balfour (1943) was prophetic with her statement that ‘the health of soil, plant, animal and man is one and indivisible.’ The Balfour's indivisibility could be better understood from the non-separation imperative since ill-health of any one domain can affect the overall health, and the connectivity hypothesis that the health of various domains is linked together (Vieweger and Doring, 2015). The environmental health assessments of chemical hazards are inadequate at present (Table 4). Though there exist different criteria and descriptors of health in each

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Possible Resistance Mechanisms Porin

Target modification Enzymatic degradation

Pesticide/Antimicrobial modifying enzymes

185

Pesticide degraders

Antibiotic/antimicrobial subsisting microorganisms

Resistant Bacterium

Overexpression of efflux pumps

Lipopolysaccharide Porin mutation mutation

Horizontal gene transfer of resistance through:

Mutation

Transformation Conjugation Transduction

Pesticides

Bacteria

Fig. 1. Pesticides as driver for antimicrobial resistance.

domain, microorganisms in all domains are significant to the health promotion or the causation of ill health. Pesticides influence human health directly or indirectly. Velmurugan et al. (2017) reported the organophosphate-induced hyperglycemia due to gluconeogenesis, mediated by the pesticide-degrading gut microbiota. Pesticides causing dysbiosis of human gut microbiome leading to diabetes necessitate extensive research whether these chemicals can trigger antimicrobial resistance. Human microbiome has many antibiotic resistance genes and any disturbance to the microbiome composition can increase the chances of horizontal gene transfer of these genes (Sommer et al., 2010). Risks are very high for the agricultural workers involved in the production, harvest, storage, transportation, and processing of food and fibres, and even the consumers who prepare contaminated food or consume raw food since several pesticides are not adequately tested for their harmful effects (Mushak and Piver, 1992; Wright et al., 2009). The microbiome of insect gut is another major interface between the animals and the environment, and can mediate resistance to insecticides (Broderick et al., 2006; Kikuchi et al., 2012; Cressey, 2013; Xia et al., 2018). The gut microbiota of honey bee, depending on their

possession of enolpyruvylshikimate-3-phosphate synthase (EPSPS) of Class I or Class II, were found to show differential response and resistance to the herbicide glyphosate, which affected the bee health due to perturbation of beneficial organisms (Motta et al., 2018). Kirubakaran et al. (2017) isolated monocrotophos-degrading Bacillus spp. (B. cereus, B. firmus and B. thuringiensis) using the minimal salt medium containing 0.4% insecticide for the enrichment, and found them to be resistant against antibiotics such as chloramphenicol (30 μg mL−1), ampicillin (10 μg mL−1), streptomycin (100 μg mL−1), cefotaxime (30 μg mL−1) and tetracycline (30 μg mL−1). The curing of plasmid in these bacterial isolates not only removed the pesticide-degrading ability Table 4 Hazard classes of the globally harmonized system of classification and labelling of chemicals. Physical hazards

Health hazards

Explosives

Acute toxicity Hazardous to aquatic (oral/dermal/inhalation) environment (acute/chronic) Skin corrosion/irritation Hazardous to the ozone layer Serious eye damage/eye irritation Respiratory or skin sensitization Germ cell mutagenicity Carcinogenicity Reproductive toxicology Target organ systemic toxicity - single exposure Target organ systemic toxicity - repeated exposure Aspiration toxicity

Flammable gases Table 3 Testing of direct effects of pesticides on microorganisms.

Aerosols

Pesticide

Concentration tested

Test organism

Reference

Carbaryl Carbofuran

9400.0 mg L−1 20.5 mg L−1

Chlorpyrifos

6600.0 mg L−1 46.3 mg L−1

E. coli O157:H7 Vibrio phosphoreum E. coli O157:H7 V. phosphoreum E. coli E. coli E. coli O157:H7

Guan et al. (2001) Somasundaram et al. (1990) Guan et al. (2001) Somasundaram et al. (1990) Harishankar et al. (2013) Higgins and Hohn (2008) Guan et al. (2001)

E. E. E. E.

Botelho et al. (2012) Staley et al. (2012) Botelho et al. (2012) Guan et al. (2001)

1400.0 mg L−1 Diazinon 0.005 μg L−1 2,4-D 31,400.0 mg L−1 230.0 mg L−1 Atrazine 102.2 mg L−1 100.0 mg L−1 Glyphosate 1,34,00.0 mg L−1 900.0 mg L−1 Chlorothalonil 169.4 mg L−1 Thiram 13,400.0 mg L−1

coli ATCC 25922 coli O157:H7 coli ATCC 25992 coli O157:H7

E. coli ATCC 25922 Botelho et al. (2012) E. coli O157:H7 Staley et al. (2012) E. coli O157:H7 Guan et al. (2001)

Oxidizing gases Gases under pressure Flammable liquids Flammable solids Self-reactive substances

Pyrophoric liquids

Pyrophoric solids Self-heating substances Substances which, in contact with water, emit flammable gases Oxidizing liquids Oxidizing solids Organic peroxides Corrosive to metals Desensitized explosives (Adapted from UN, 2017).

Environmental hazards

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but also antibiotic resistance. In-silico analysis revealed that the enzyme, organophosphorus hydrolase, involved in the degradation of monocrotophos could bind to antibiotics (Kirubakaran et al., 2017). Future experimentation is required to assess the hydrolytic activity of organophosphorus hydrolase due to binding with antibiotics, and the alterations in the capabilities for pesticide degradation and the development of antibiotic resistance. Dada et al. (2018) reported changes in the composition of microbiota and its functions between fenitriothion resistant (FEN_Res) and susceptible (FEN_Sus) populations of a species of mosquito, Anopheles albimanus. Since pesticides can induce multidrug resistance (Kirubakaran et al., 2017, 2018), the enrichment of microbiota with pesticide-antibiotic cross-resistance can led to the evolution of new resistance profiles for pesticides in insects and for antimicrobials in the animals including humans with insect infestation. In the case of plant microbiome, ‘Human pathogens on plants’ (HPOPs) which can enter and colonize the plant tissues and interact with the plants can also be a component of great concern (Fletcher et al., 2013). These HPOPs include species of Escherichia, Salmonella and Shigella of the bacterial family Enterobacteriaceae; other genera such as Enterobacter, Erwinia, Pantoea and Pectobacterium of this family are plant pathogens. Both the HPOPs and plant pathogens have many common genes that contribute to virulence in plants and vertebrates (Fletcher et al., 2013; Ximenes et al., 2017). Pesticides may disturb the plant microbiomes, especially those of endophytic in nature by systemic chemicals and the epiphytes by contact chemicals and encourage the horizontal transfer of antimicrobial resistance between the HPOPS and the plant pathogens. Such developments can alter the ecological cycle of human enteric pathogens in plants, animals and humans, more so when plants can transmit bacteria with multidrug (antimicrobial) resistance vertically through seeds. The holistic perspective on antibiotic resistance among all the domains, envisaged in the ‘One Health’ approach is currently lacking (Atlas and Maloy, 2014). Soils are an important source of natural antibiotics and antibiotic resistance as well (Brevik and Sauer, 2015). Since the stress conditions influence the production of antibiotic-like compounds, the contamination of soils due to the pesticides, organic chemicals or antibiotics can increase the incidence and persistence of antibiotic resistant genes (Swiecilo and Zych-Wezyk, 2013; Steffan et al., 2018). Aarestrup (2015) recognized the livestock reservoir for antimicrobial resistance having impact on human health. Watts et al. (2017) considered that the aquaculture systems and farms, with the combined use of pesticides and antibiotics, could be the ‘genetic reactors’ or hot spots for antimicrobial resistance genes. Woappi et al. (2016) even proposed the term ‘antibiotrophs’ for those capable of subsisting in environments with elevated concentrations of antibiotics or with the use of these chemicals as sole sources of carbon. Rapid and reliable technologies for risk assessment and prevention of antimicrobial resistance, particularly when the stressors are pesticides, are critical to the success of ‘One Health’ approach. 6. Conclusions The burden of disease due to human exposure to toxic chemicals has been estimated (WHO, 2016). The pesticide self-poisoning is a major contributor to the global burden of human suicides. There were approximately 110,000 deaths every year due to the pesticide self-poisoning between the years 2010 and 2014 (Mew et al., 2017). But, the human and environmental consequences of pesticides, especially due to the alterations of functional traits and processes mediated by the soil or foliar microbiome, remain unfathomable. The contaminated food, feed, and fodder, not only with pesticides but also with the pesticide-degrading and/or antibiotic or antimicrobial subsisting microorganisms, may contribute indirectly to the poor health of animals and humans. Zhang et al. (2017) showed that the loss of microbial diversity in soils led to increased uptake of insecticides (acetamiprid and imidacloprid) by the host plants Brassica. Even the plant-derived compounds such as ascorbic

acid are found to induce degradation of pesticides (Hou et al., 2017). The pesticides in general and antimicrobials, when used in agriculture and in the public health, may cause a combined selection pressure and coselection for tolerance, resistance, persistence or degradation. Hence, these chemicals might impact the dynamics of pest population, and the diversity and functioning of microorganisms which are of high ecological economic value, probably more than the costs involved in their discovery and marketing. Intrinsic reactivity of pesticides and even the possible modes of degradation can be predicted from the molecular structures but the quantitative prediction of their action on the pests or their degradation by microorganisms in the field remains inadequate (Fenner et al., 2013). Recently, the WHO (2017) identified 12 bacteria as the greatest threat to human health because of their resistance to antibiotics. The challenge now is to gain further insights on the contributions of human activities such as the agricultural and public health application of pesticides on the global spread of these antimicrobial resistant bacteria of critical, high, and medium priorities and on the evolution of newer members. The environmental release of newly synthesized pesticidal chemicals and the genetically engineered or synthetic microorganisms created with intent of producing new drugs might influence the population dynamics and functioning of microbiome in soils or foliar regions significantly. The predictions about the environmental consequences of the application of pesticides and antimicrobials continue to be difficult to make even with increasing knowledge on the chemistry of synthesized chemical substances.

Acknowledgements BR and KV are grateful to the Government of Australia (Department of Education, Employment and Workplace Relations) for the Endeavour Research Fellowship and Endeavour Executive Award, respectively. References Aarestrup, F.M., 2015. The livestock reservoir for antimicrobial resistance: a personal view of changing patterns of risks, effects of interventions and the way forward. Philos. Trans. R. Soc. Lond. Ser. B Biol. Sci. 370 (1670), 20140085. Achouak, W., Heulin, T., Pages, J.-M., 2001. Multiple facets of bacterial porins. FEMS Microbiol. Lett. 199, 1–7. Alcott, B., 2005. Jevon's paradox. Ecol. Econ. 54, 9–21. Alekshun, M.N., Levy, S.B., 1999. The mar regulon: multiple resistance to antibiotics and other toxic chemicals. Trends Microbiol. 7, 410–413. Alexander, M., 1982. Aging, bioavailability, and overestimation of risk from environmental pollutants. Environ. Sci. Technol. 34, 4259–4265. Alexandre, G., Greer, S.E., Zhulin, I.B., 2000. Energy taxis is a dominant behaviour in Azospirillum brasilense. J. Bacteriol. 182, 6042–6048. Amabile-Cuevas, C.F., 2013. Antibiotic resistance: from Darwin to Lederberg to Keynes. Microb. Drug Resist. 19, 73–87. Aminov, R.I., 2013. Biotic acts of antibiotics. Front. Microbiol. 4, 241. Andrews, J.M., 2001. Determination of minimum inhibitory concentrations. J. Antimicrob. Chemother. 48 (Suppl. S1), 5–16. Arias-Estevez, M., Lopez-Periago, E., Martinez-Carballo, E., Simal-Gandara, J., Mejuto, J.-C., Garcia-Rio, L., 2008. The mobility and degradation of pesticides in soils and the pollution of groundwater resources. Agric. Ecosyst. Environ. 123, 247–260. Atlas, R.M., Maloy, S., 2014. The future of One Health. Microbiol. Spectrum 2 (1) (OH0018-2012). Atwood, D., Paisley-Jones, C., 2017. Pesticides industry sales and usage. 2008–2012 Market estimates. Office of Chemical Safety and Pollution Prevention, U. S. Environmental Protection Agency, Washington, DC 20460. https://www.epa.gov/sites/production/ files/2017-01/documents/pesticides-industry-sales-usage-2016_0.pdf. Audus, L.J., 1949. The biological detoxication of 2, 4-dichlorophenoxyacetic acid in soil. Plant Soil 2, 31–36. Bachmann, A., de Bruin, W., Jumelet, J.C., Rijnaarts, H.H.N., Zehnder, A.J.B., 1988. Aerobic biomineralization of alpha-hexachlorocyclohexane in contaminated soil. Appl. Environ. Microbiol. 54, 548–554. Balfour, E.B., 1943. The Living Soil. Faber and Faber, London. Baltz, R.H., 2007. Antimicrobials from actinomycetes: Back to the future. Microbe 2, 125–131. Barriuso, E., Benoit, P., Dubus, I.G., 2008. Formation of pesticide nonextractable (bound) residues in soil: magnitude, controlling factors and reversibility. Environ. Sci. Technol. 42, 1845–1854. Begic, S., Worobec, E.A., 2006. Regulation of Serratia marcescens ompE and ompC porin genes in response to osmotic stress, salicylate, temperature and pH. Microbiology 152, 485–491.

B. Ramakrishnan et al. / Science of the Total Environment 654 (2019) 177–189 Benbrook, C.M., 2016. Trends in glyphosate herbicide use in the United States and globally. Environ. Sci. Eur. 28, 3. Bergman, J., 2003. Does the acquisition of antibiotic and pesticide resistance provide evidence for evolution? J. Creation 17, 26–32. Bhuyan, S., Sreedharan, B., Adhya, T.K., Sethunathan, N., 1993. Enhanced degradation of γhexachlorocyclohexane (γ-HCH) in HCH (commercial) acclimatized flooded soil: factors affecting its development and persistence. Pestic. Sci. 38, 49–55. Bigley, A.N., Raushel, F.M., 2013. Catalytic mechanisms for phosphotriesterases. Biochim. Biophys. Acta 1834, 443–453. Blanco, P., Hernando-Amado, S., Reales-Calderon, J.A., Corona, F., Lira, F., Alcalde-Rico, M., Bernardini, A., Sanchez, M.B., Martinez, J.L., 2016. Bacterial multidrug efflux pumps: much more than antibiotic resistance determinants. Microorganisms 4, 14. Boesten, J.J.T.I., 2016. Proposal for field-based definition of soil bound pesticide residues. Sci. Total Environ. 544, 114–117. Borgesen, C.D., Fomsgaard, I.S., Plauborg, F., Schelde, K., Spliid, N.H., 2015. Fate of Pesticides in Agricultural Soils. DCA Report No. 062. Nationalt center for fodevarer og jordbrug, Aarhus University, Tjele 978-87-93176-76-8, p. 73. Botelho, R.G., Froes, C.M., Santos, J.B., 2012. Toxicity of herbicides on Escherichia coli growth. Braz. J. Biol. 72, 141–146. Brauner, A., Fridman, O., Gefen, O., Balaban, N.Q., 2016. Distinguishing between resistance, tolerance and persistence to antibiotic treatment. Nat. Rev. Microbiol. 14, 320–330. Bren, A., Eisenbach, M., 2000. How signals are heard during bacterial chemotaxis: proteinprotein interactions in sensory signal propagation. J. Bacteriol. 182, 6865–6873. Brevik, E.C., Sauer, T.J., 2015. The past, present and future of soils and human health studies. Soil 1, 35–36. Broderick, N.A., Raffa, K.F., Handelsman, J., 2006. Midgut bacteria required for Bacillus thuringiensis insecticidal activity. Proc. Natl. Acad. Sci. U. S. A. 103, 15196–15199. Bromilow, R.H., 1999. Book review of “Pesticide bound residues in soil”. Workshop – Senate Commission for the Assessment of Chemicals used in agriculture, Report 2. WileyVCH Verlag GmbH, Weinheim, Germany 1998. pp 186. Pestic. Sci. 55, 755–756. Bugg, T.D.H., Ramaswamy, S., 2008. Non-heme iron-dependent dioxygenases: unravelling catalytic mechanisms for complex enzymatic oxidations. Curr. Opin. Chem. Biol. 12, 134–140. Bush, K., Jacoby, G.A., 2010. Updated functional classification of β-lactamases. Antimicrob. Agents Chemother. 54, 969–976. Carson, R., 1962. Silent Spring. Houghton Mifflin, Boston (272 p). CDC, 2006. Centers for disease control and prevention, emergence of mycobacterium tuberculosis with extensive resistance to second-line drugs-worldwide, 2000–2004. MMWR Morb. Mortal. Wkly Rep. 55, 301–305. Chakrabarty, A.M., 1972. Genetic basis of the biodegradation of salicylate in Pseudomonas. J. Bacteriol. 112, 815–823. Cohen, S.P., Levy, S.B., Foulds, J., Rosner, J.L., 1993. Salicylate induction of antibiotic resistance in Escherichia coli: activation of the mar operon and a mar-independent pathway. J. Bacteriol. 175, 7856–7862. Copley, S.D., 2012. Toward a systems biology perspective on enzyme evolution. J. Biol. Chem. 287, 3–10. Copley, S.D., Rokicki, J., Turner, P., Daligault, H., Nolan, M., Land, M., 2012. The whole genome sequence of Sphingobium chlorophenolicum L-1: insights into the evolution of the pentachlorophenol degradation pathway. Genome Biol. Evol. 4, 184–198. Cressey, D., 2013. Europe debates risk to bees. Nature 496, 408. Cribb, J., 2017. Surviving the 21st Century. Humanity's Ten Great Challenges and how we Can Overcome Them. Springer 978-3-319-412702. CropLife International, 2017. Obsolete and unwanted pesticide stocks. Practical guidance on safeguarding, disposal and prevention. www.croplife.org (Brussels, Belgium. accessed on July 20, 2018). Cuesta, S.M., Rahman, S.A., Furnham, N., Thornton, J.M., 2015. The classification and evolution of enzyme function. Biophys. J. 109, 1082–1086. Dada, N., Sheth, M., Liebman, K., Pinto, J., Lenhart, A., 2018. Whole metagenome sequencing reveals links between mosquito microbiota and insecticide resistance in malaria vectors. Sci. Rep. 8, 2084. Dantas, G., Sommer, M.O.A., Oluwasegun, R.D., Church, G.M., 2008. Bacteria subsisting on antibiotics. Science 320, 100–103. D'Costa, V.M., McGrann, K.M., Hughes, D.W., Wright, G.D., 2006. Sampling the antibiotic resistome. Science 311, 374–377. Delcour, I., Spanoghe, P., Uyttendaele, M., 2015. Literature review: impact of climate change on pesticide use. Food Res. Int. 68, 7–15. DiGiovanni, G.D., Neilson, J.W., Pepper, I.L., Sinclair, N.A., 1996. Gene transfer of Alcaligenes eutrophus JMP134 plasmid pJP4 to indigenous soil recipients. Appl. Environ. Microbiol. 62, 2521–2526. Don, R.H., Pemberton, J.M., 1981. Properties of six pesticide degradation plasmids isolated from Alcaligenes paradoxus and Alcaligenes eutrophus. J. Bacteriol. 145, 681–686. Don, R.H., Pemberton, J.M., 1985. Genetic and physical map of the 2,4dichlorophenoxyacetic acid-degradative plasmid pJP4. J. Bacteriol. 161, 466–468. Dungan, R.S., McKinney, C.W., Leytem, A.B., 2018. Tracking antibiotic resistance genes in soil irrigated with diary wastewater. Sci. Total Environ. 635, 1477–1483. Dzidic, S., Suskovic, J., Kos, B., 2008. Antibiotic resistance mechanisms in bacteria: biochemical and genetic aspects. Food Technol. Biotechnol. 46, 11–21. ECHA, 2017. Guidance on information requirements and chemical safety assessment. Chapter R. 7b: Endpoint specific guidance. Version 4.0. European Chemicals Agency, FI-00121 Helsinki, Finland. Egli, T., 2010. How to live at very low substrate concentration. Water Res. 44, 4826–4837. Epand, R.M., Walker, C., Epand, R.F., Magarvey, N.A., 2016. Molecular mechanisms of membrane targeting antibiotics. Biochim. Biophys. Acta 1858, 980–987. FAO, 2017. Prevention and disposal of obsolete pesticides. Food and Agriculture Organization of the United Nations, Pest and Pesticide Management Team, Plant Production and Protection Division, Rome, Italywww.fao.org/agriculture/crops/obsolete-pesticides/prevention...obsolete-pesticides/en/, Accessed date: 20 July 2018.

187

Fenner, K., Canonica, S., Wackett, L.P., Elsner, M., 2013. Evaluating pesticide degradation in the environment: blind spots and emerging opportunities. Science 341, 752–758. Fernandez, L., Hancock, R.E., 2012. Adaptive and mutational resistance: role of porins and efflux pumps in drug resistance. Clin. Microbiol. Rev. 25, 661–681. Fischer, H., Ingwersen, J., Kuzyakov, Y., 2010. Microbial uptake of low-molecular-weight organic substances out-competes sorption in soil. Eur. J. Soil Sci. 61, 504–513. Fisher, P.R., Appleton, J., Pemberton, J.M., 1978. Isolation and characterization of the pesticide-degrading plasmid pJP1 from Alcaligenes paradoxus. J. Bacteriol. 135, 798–804. Fletcher, J., Leach, J.E., Eversole, K., Tauxe, R., 2013. Human pathogens on plants: designing a multidisciplinary strategy for research. Phytopathology 103, 306–315. Forsberg, K.J., Reyes, A., Wang, B., Selleck, E.M., Sommer, M.O.A., Dantas, G., 2012. The shared antibiotic resistome of soil bacteria and human pathogens. Science 337, 1107–1111. Forsberg, K.J., Patel, S., Gibson, M.K., Lauber, C.L., Knight, R., Fierer, N., Dantas, G., 2014. Bacterial phylogeny structures soil resistomes across habitats. Nature 509, 612–616. Fox, J.L., 1983. Soil microbes pose problems for pesticides. Science 221, 1029–1031. Fraise, A.P., 2002. Biocide abuse and antimicrobial resistance – a cause for concern? J. Antimicrob. Chemother. 49, 11–12. https://doi.org/10.1093/jac/49.1.11. Franklin, A.M., Aga, D.S., Cytryn, E., Durso, L.M., McLain, J.E., Pruden, A., Roberts, M.C., Rothrock Jr., M.J., Snow, D.D., Watson, J.E., Dungan, R.S., 2016. Antibiotics in agroecosystems: introduction to the special section. J. Environ. Qual. 45, 377–393. Furlan, L., Pozzebon, A., Duso, C., Simon-Delso, N., Sanchez-Bayo, F., Marchand, P.A., Codato, F., van Lexmond, M.B., Bonmatin, J.-M., 2018. An update of the worldwide integrated assessment (WIA) on systemic insecticides. Part 3: alternatives to systemic insecticides. Environ. Sci. Pollut. Res. https://doi.org/10.1007/s11356-017-1052-5. Galperin, M.Y., Koonin, E.V., 2012. Divergence and convergence in enzyme evolution. J. Biol. Chem. 287, 21–28. Gan, J., Wang, Q., Yates, S.R., Koskinen, W.C., Jury, W.A., 2002. Dechlorination of chloroacetanilide herbicides by thiosulfate salts. Proc. Natl. Acad. Sci. U. S. A. 99, 5189–5194. Gao, J., Ellis, L.B., Wackett, L.P., 2010. The University of Minnesota biocatalysis/biodegradation database: improving public access. Nucleic Acids Res. 38, D488–D491. Gefen, O., Balaban, N.Q., 2009. The importance of being persistent: heterogeneity of bacterial populations under antibiotic stress. FEMS Microbiol. Rev. 33, 704–717. Gevao, B., Semple, K.T., Jones, K.C., 2000. Bound pesticide residues in soils: a review. Environ. Pollut. 108, 3–14. Grimsley, J.K., Scholtz, J.M., Pace, C.N., Wild, J.R., 1997. Organophosphorus hydrolase is a remarkably stable enzyme that unfolds through a homodimeric intermediate. Biochemistry 36, 14366–14374. Guan, T.Y., Blank, G., Ismond, A., Van Acker, R., 2001. Fate of foodborne bacterial pathogens in pesticide products. J. Sci. Food Agric. 81, 503–512. Guillette Jr., L.J., Iguchi, T., 2012. Life in a contaminated world. Science 337, 1614–1615. Hall, B.G., Barlow, M., 2004. Evolution of the serine β-lactamases: past, present and future. Drug Resist. Updat. 7, 111–123. Harishankar, M.K., Sasikala, C., Ramya, M., 2013. Efficiency of the intestinal bacteria in the degradation of the toxic pesticide, chlorpyrifos. 3 Biotech 3, 137–142. Harper, L.L., McDaniel, C.S., Miller, C.E., Wild, J.R., 1988. Dissimilar plasmids isolated from Pseudomonas diminuta MG and a Flavobacterium sp. (ATCC 27551) contain identical opd genes. Appl. Environ. Microbiol. 54, 2586–2589. Hawkins, A.C., Harwood, C.S., 2002. Chemotaxis of Ralstonia eutropha JMP134 (pJP4) to the herbicide, 2,4-dichlorophenoxyacetate. Appl. Environ. Microbiol. 68, 968–972. Head, I.M., 1998. Bioremediation: towards a credible technology. Microbiology 144, 599–608. Higgins, J., Hohn, C., 2008. Effects of prevalent freshwater chemical contaminants on in vitro growth of Escherichia coli and Klebsiella pneumoniae. Environ. Pollut. 152, 259–266. Hildebrandt, A., Lacorte, S., Barcelo, D., 2009. Occurrence and fate of organochlorinated pesticides and PAH in agricultural soils from the Ebro River Basin. Arch. Environ. Contam. Toxicol. 57, 247–255. Hill, I.R., Wright, S.J.L., 1978. Pesticide Microbiology: Microbiological Aspects of Pesticide Behaviour in the Environment. Academic Press, London 9780123486509. Horne, I., Sutherland, T.D., Harcourt, R.L., Russell, R.J., Oakeshott, J.G., 2002. Identification of an opd (organophosphate degradation) gene in an Agrobacterium isolate. Appl. Environ. Microbiol. 68, 3371–3376. Hou, X., Huang, X., Ai, Z., Zhao, J., Zhang, L., 2017. Ascorbic acid induced atrazine degradation. J. Hazard. Mater. 327, 71–78. Hou, X., Huang, X., Li, M., Zhang, Y., Yuan, S., Ai, Z., Zhao, J., Zhang, L., 2018. Fenton oxidation of organic contaminants with aquifer sediment activated by ascorbic acid. Chem. Eng. J. 348, 255–262. Huang, X., He, J., Yan, X., Hong, Q., Chen, K., He, Q., Zhang, L., Liu, X., Chuang, S., Li, S., Jiang, J., 2017. Microbial catabolism of chemical herbicides: microbial resources, metabolic pathways and catabolic genes. Pestic. Biochem. Physiol. 143, 272–297. Hussain, S., Hartley, C.J., Shettigar, M., Pandey, G., 2016. Bacterial biodegradation of neonicotinoid pesticides in soil and water systems. FEMS Microbiol. Lett. 363 (23), fnw252. IUPAC, 1993. Glossary for chemists of terms used in toxicology (IUPAC Recommendations 1993): 2020. The International Union of Pure and Applied Chemistry (IUPAC)https:// goldbook.iupac.org/html/B/B00656.html, Accessed date: 20 July 2018. Kaminski, M.A., Furmanczyk, E.M., Sobczak, A., Dziembowski, A., Lipinski, L., 2017. Pseudomonas silesiensis sp. nov. strain A3T isolated from a biological pesticide sewage treatment plant and analysis of the complete genome sequence. Syst. Appl. Microbiol. 41, 13–22. Kawahara, K., Tanaka, A., Yoon, J., Yokota, A., 2010. Reclassification of a parathionedegrading Flavobacterium sp. ATCC 27551 as Sphingobium fuliginis. J. Gen. Appl. Microbiol. 56, 249–255.

188

B. Ramakrishnan et al. / Science of the Total Environment 654 (2019) 177–189

Kemper, N., 2008. Veterinary antibiotics in the aquatic and terrestrial environment. Ecol. Indic. 8, 1–13. Kennedy, I.R., Skerritt, J.H., Johnson, G.I., Highley, E., 1998. Seeking agricultural produce free of pesticide residues. Proceedings of an International workshop held in Yogyakarta, Indonesia, 17–19 February 1998. ACIAR Proceedings No. 85, 406 p. Arawang Communication Group, Canberra, Australia. Kikuchi, Y., Hayatsu, M., Hosokawa, T., Nagayama, A., Tago, K., Fukatsu, T., 2012. Symbiont-mediated insecticide resistance. Proc. Natl. Acad. Sci. U. S. A. 109, 8618–8622. Kirubakaran, R., Murugan, A., Natarajan, D., Parray, J., 2017. Emergence of multi drug resistance among soil bacteria exposing to insecticides. Microb. Pathog. 105, 153–165. Kirubakaran, R., Murugan, A., Natarajan, D., Gopinath, S., Parray, J.A., Aruljothi, K.N., Shameem, N., Alqarawi, A.A., Hashem, A., Abd_Allah, E.F., 2018. Pesticide degrading natural multidrug resistance bacterial flora. Microb. Pathog. 114, 304–310. Krell, T., Lacal, J., Munoz-Martinez, F., Reyes-Darias, J.A., Cadirci, B.H., Garcia-Fontana, C., Ramos, J.L., 2011. Diversity at its best: bacterial taxis. Environ. Microbiol. 13, 1115–1124. Kuppusamy, S., Kakarla, D., Venkateswarlu, K., Megharaj, M., Yoon, Y.-E., Lee, Y.B., 2018. Veterinary antibiotics (VAs) contamination is a global agro-ecological issue: a critical view. Agric. Ecosyst. Environ. 257, 47–59. Kurenbach, B., Marjoshi, D., Amabile-Cuevas, C.F., Ferguson, G.C., Godsoe, W., Gibson, P., Heinemann, J.A., 2015. Sublethal exposure to commercial formulations of the herbicides dicamba, 2,4-dichlorophenoxyacetic acid, and glyphosate cause changes in antibiotic susceptibility in Escherichia coli and Salmonella enterica serovar Typhimurium. MBio 6, e00009–e00015. Lacal, J., Garcia-Fontana, C., Munoz-Martinez, F., Ramos, J.-L., Krell, T., 2010. Sensing of environmental signals: classification of chemoreceptors according to the size of their ligand binding regions. Environ. Microbiol. 12, 2873–2884. Lacal, J., Reyes-Darias, J.A., Garcia-Fontana, C., Ramos, J.-L., Krell, T., 2012. Tactic responses to pollutants and their potential to increase biodegradation efficiency. J. Appl. Microbiol. 114, 923–933. Lamberth, C., Jeanmart, S., Luksch, T., Plant, A., 2013. Current challenges and trends in the discovery of agrochemicals. Science 341, 742–746. Lau, C.H.-F., van Engelen, K., Gordon, S., Renaud, J., Topp, E., 2017. Novel antibiotic resistance determinants from agricultural soil exposed to antibiotics widely used in human medicine and animal farming. Appl. Environ. Microbiol. 83, e00989-17. Levy, S.B., 1992. Active efflux mechanisms for antimicrobial resistance. Antimicrob. Agents Chemother. 36, 695–703. Lewis, K.A., Tzilivakis, J., Warner, D.J., Green, A., 2016. An international database for pesticide risk assessments and management. Hum. Ecol. Risk Assess. Int. J. 22 (4), 1050–1064. Li, X.-Z., Nikaido, H., 2004. Efflux-mediated drug resistance in bacteria. Drugs 64, 159–204. Li, X.Z., Poole, K., 1999. Organic solvent-tolerant mutants of Pseudomonas aeruginosa display multiple antibiotic resistance. Can. J. Microbiol. 45, 18–22. Liu, X., Parales, R.E., 2009. Bacterial chemotaxis to atrazine and related s-triazines. Appl. Environ. Microbiol. 75, 5481–5488. Liu, L.K., Dai, Y., Abdelwahab, H., Sobrado, P., Tanner, J.J., 2018. Structural evidence for rifampicin monooxygenase inactivating rifampicin by cleaving its ansa-bridge. Biochemistry 57, 2065–2068. Lorca, G.L., Barabote, R.D., Zlotopolski, V., Tran, C., Winnen, B., Hvorup, R.N., Stonestrom, A.J., Nguyen, E., Huang, L.-W., Kim, D.S., Saier Jr., M.H., 2007. Transport capabilities of eleven gram-positive bacteria: comparative genomic analyses. Biochim. Biophys. Acta 1768, 1342–1366. Luo, Y.Z., Zhang, M.H., 2010. Spatially distributed pesticide exposure assessment in the Central Valley, California, USA. Environ. Pollut. 158, 1629–1637. MacRae, I.C., Raghu, K., Castro, T.F., 1967. Persistence and biodegradation of four common isomers of benzene hexachloride in submerged soils. J. Agric. Food Chem. 15, 911–914. Markley, J.L., Wencewicz, T.A., 2018. Tetracycline-inactivation enzymes. Front. Microbiol. 9, 1058. Martini, M.C., Wibberg, D., Lozano, M., Tejerizo, G.T., Albicoro, F.J., Jaenicke, S., van Elsas, J.D., Petroni, A., Garcillan-Barcia, M.P., de la Cruz, F., Schluter, A., Puhler, A., Pistorio, M., Lagares, A., Del Papa, M.F., 2016. Genomics of high molecular weight plasmids isolated from an on-farm biopurification system. Sci. Rep. 6, 28284. Matheson, V.G., Forney, L.J., Suwa, Y., Nakatsu, C.H., Sexstone, A.J., Holben, W.E., 1996. Evidence for acquisition in nature of a chromosomal 2,4-dichlorophenoxyacetic acid/αketoglutarate dioxygenase gene by different Burkholderia spp. Appl. Environ. Microbiol. 62, 2457–2463. McDougall, P., 2016. The cost of new agrochemical product discovery, development and registration in 1995, 2000, 2005–8 and 2010 to 2014. R&D expenditure in 2014 and expectations for 2019. A consultancy study for CropLife International, CropLife America and the European Crop Protection Association. p. 41. https://www. croplifeamerica.org/wp.../Phillips-McDougall-Final-Report_4.6.16.pdf. Megharaj, M., Singleton, I., Kookana, R., Naidu, R., 1999. Persistence and effects of fenamiphos on native algal populations and enzymatic activities in soil. Soil Biol. Biochem. 31, 1549–1553. Megharaj, M., Ramakrishnan, B., Venkateswarlu, K., Sethunathan, N., Naidu, R., 2011. Bioremediation approaches for organic pollutants: a critical perspective. Environ. Int. 37, 1362–1375. Menon, P., Gopal, M., Parsad, R., 2005. Effects of chlorpyrifos and quinalphos on dehydrogenase activities and reduction of Fe3+ in the soils of two semi-arid fields of tropical India. Agric. Ecosyst. Environ. 108, 73–83. Mew, E.J., Padmanathan, P., Konradsen, F., Eddleston, M., Chang, S.-S., Phillips, M.R., Gunnell, D., 2017. The global burden of fatal self-poisoning with pesticides 2006–15: systematic review. J. Affect. Disord. 219, 93–104.

Michael, A.J., 2017. Evolution of biosynthetic diversity. Biochem. J. 474, 2277–2299. Monard, C., Martin-Laurent, F., Devers-Lamrani, M., Lima, O., Vandenkoornhuyse, P., Binet, F., 2010. atz gene expressions during atrazine degradation in the soil drilosphere. Mol. Ecol. 19, 749–759. Morar, M., Pengelly, K., Koteva, K., Wright, C.D., 2012. Mechanism and diversity of the erythromycin esterase family of enzymes. Biochemistry 51, 1740–1751. Motta, E.V.S., Raymann, K., Moran, N.A., 2018. Glyphosate perturbs the gut microbiota of honey bees. Proc. Natl. Acad. Sci. U. S. A. https://doi.org/10.1073/pnas.1803880115. Mukhtar, T.A., Koteva, K.P., Hughes, D.W., Wright, G.D., 2001. Vgb from Staphylococcus aureus inactivates streptogramin B antibiotics by an elimination mechanism not hydrolysis. Biochemistry 40, 8877–8886. Mulbry, W.W., Karns, J.S., 1989. Parathion hydrolase specified by the Flavobacterium opd gene: relationship between the gene and protein. J. Bacteriol. 171, 6740–6746. Mushak, E.W., Piver, W.T., 1992. Agricultural chemical utilization and human health. Environ. Health Perspect. 97, 269–274. Newby, D.T., Josephson, K.L., Pepper, I.L., 2000a. Detection and characterization of plasmid pJP4 transfer to indigenous soil bacteria. Appl. Environ. Microbiol. 66, 290–296. Newby, D.T., Gentry, T.J., Pepper, I.L., 2000b. Comparison of 2,4-dichlorophenoxyacetic acid degradation and plasmid transfer in soil resulting from bioaugmentation with two different pJP4 donors. Appl. Environ. Microbiol. 66, 3399–3407. Nielsen, T.K., Rasmussen, M., Demaneche, S., Cecillon, S., Vogel, T.M., Hansen, L.H., 2017. Evolution of Sphingomonad gene clusters related to pesticide catabolism revealed by genome sequence and mobilomics of Sphingobium herbicidovorans MH. Genome Biol. Evol. 9, 2477–2490. Nordenholt, R.M., Goyne, K.W., Kremer, R.J., Lin, C.-H., Lerch, R.N., Veum, K.S., 2016. Veterinary antibiotic effects on atrazine degradation and soil microorganisms. 45, 565–575. OECD, 2014. Report of an OECD survey on the assessment of the risks from obsolete pesticides in OECD countries. Series on Pesticides No. 79. Organisation for Economic Cooperation and Development, Pariswww.oecd.org/chemicalsafety, Accessed date: 20 July 2018. Ortiz-Hernandez, Ma.L., Sanchez-Salinas, E., Dantan-Gonzalez, E., Castrejon-Godinez, M.L., 2013. Pesticide biodegradation: mechanisms, genetics and strategies to enhance the process. In: Chamy, R., Rosenkranz, F. (Eds.), Biodegradation-Life of Science. InTech https://doi.org/10.5772/56098 Available from. https://www.intechopen.com/books/ biodegradation-life-of-science/pesticide-biodegradation-mechanisms-genetics-andstrategies-to-enhance-the-process. Pandey, G., Jain, R.K., 2002. Bacterial chemotaxis toward environmental pollutants: role in bioremediation. Appl. Environ. Microbiol. 68, 5789–5795. Parales, R.E., Ditty, J.L., Harwood, C.S., 2000. Toluene-degrading bacteria are chemotactic towards the environmental pollutants benzene, toluene, and trichloroethylene. Appl. Environ. Microbiol. 66, 4098–4104. Parales, R.E., Luu, R.A., Hughes, J.G., Ditty, J.L., 2015. Bacterial chemotaxis to xenobiotic chemicals and naturally-occurring analogs. Curr. Opin. Biotechnol. 33, 318–326. Pathak, R.K., Dikshit, A.K., 2012. Atrazine and its use. Int. J. Res. Chem. Environ. 2, 1–6. Paulsen, I.T., Brown, M.H., Skurray, R.A., 1998a. Characterization of the earliest known Staphylococcus aureus plasmid encoding a multidrug efflux system. J. Bacteriol. 180, 3477–3479. Paulsen, I.T., Sliwinski, M.K., Saier, M.H., 1998b. Microbial genome analyses: global comparisons of transport capabilities based on phylogenies, bioenergetics and substrate specificities. J. Mol. Biol. 277, 573–592. Pazos, F., Guijas, D., Valencia, A., De Lorenzo, V., 2005. MetaRouter: bioinformatics for bioremediation. Nucleic Acids Res. 33, D588–D592. Pemberton, J.M., Fisher, P.R., 1977. 2,4-D plasmids and persistence. Nature 268, 732–733. Penuelas, J., Terradas, J., 2014. The foliar microbiome. Trends Plant Sci. 19, 278–280. Pesticide Action Network Europe, 2010. Environmental Effects of Pesticides. An impression of recent scientific literature https://www.pan-europe.info, Accessed date: 20 July 2018. Pimentel, D., 1995. Amounts of pesticides reaching target pests: environmental impacts and ethics. J. Agric. Environ. Ethics 8, 17–29. Pretty, J., Bharucha, Z.P., 2015. Integrated pest management for sustainable intensification of agriculture in Asia and Africa. Insects 6, 152–182. Reddy, B.R., Sethunathan, N., 1983. Mineralization of parathion in the rice rhizosphere. Appl. Environ. Microbiol. 45, 826–829. Rivera-Ramirez, X.O., Hernandez-Ojeda, S.L., Espinosa-Aguirre, J.J., Camacho-Carranza, R., 2016. Antibiotic resistance mutations in bacteria selected by exposure to pesticides. XIV International Congress of Toxicology. Toxicol. Lett. 259, S218–S219. Rosner, J.L., 1985. Nonheritable resistance to chloramphenicol and other antibiotics induced by salicylates and other chemotactic repellents in Escherichia coli K-12. Proc. Natl. Acad. Sci. U. S. A. 82, 8771–8774. Russell, R.J., Scott, C., Jackson, C.J., Pandey, R., Pandey, G., Taylor, M.C., Coppin, C.W., Liu, J.-W., Oakeshott, J.G., 2011. The evolution of new enzyme function: lessons from xenobiotic metabolizing bacteria versus insecticide-resistant insects. Evol. Appl. 4, 225–248. Ruuskanen, M., Muurinen, J., Meierjohan, A., Parnanen, K., Tamminen, M., Lyra, C., Kronberg, L., Virta, M., 2016. Fertilizing with animal manure disseminates antibiotic resistance genes to the farm environment. J. Environ. Qual. 45, 488–493. Sabatier, P., Poulenard, J., Fanget, B., Reyss, J.-L., Develle, A.-L., Wilhelm, B., Ployon, E., Pignol, C., Naffrechoux, E., Dorioz, J.-M., Montuelle, B., Arnaud, F., 2014. Long-term relationships among pesticide applications, mobility, and soil erosion in a vineyard watershed. Proc. Natl. Acad. Sci. U. S. A. 111, 15647–15652. Sahu, S.K., Patnaik, K.K., Sharmila, M., Sethunathan, N., 1990. Degradation of alpha-, beta-, and gamma-hexachlorocyclohexane by a soil bacterium under aerobic conditions. Appl. Environ. Microbiol. 56, 3620–3622. Sakai, Y., Ogawa, N., Shimomura, Y., Fujii, T., 2014. A 2,4-dichlorophenoxyacetic acid degradation plasmid pM7012 discloses distribution of an unclassified megaplasmid group across bacterial species. Microbiology 160, 525–536.

B. Ramakrishnan et al. / Science of the Total Environment 654 (2019) 177–189 Samanta, S.K., Bhushan, B., Chauhan, A., Jain, R.K., 2000. Chemotaxis of a Ralstonia sp. SJ98 toward different nitroaromatic compounds and their degradation. Biochem. Biophys. Res. Commun. 269, 117–123. Schreinemachers, P., Tipraqsa, P., 2012. Agricultural pesticides and land use intensification in high, middle and low-income countries. Food Policy 37, 616–626. Schwarzenbach, R.P., Escher, B.I., Fenner, K., Hofstetter, T.B., Johnson, C.A., von Gunten, U., Wehrli, B., 2006. The challenge of micropollutants in aquatic systems. Science 313, 1072–1077. Scott, C., Pandey, G., Hartley, C.J., Jackson, C.J., Cheesman, M.J., Taylor, M.C., Pandey, R., Khurana, J.L., Teese, M., Coppin, C.W., Weir, K.M., Jain, R.K., Lal, R., Russell, R.J., Oakeshott, J.G., 2008. The enzymatic basis for pesticide bioremediation. Indian J. Microbiol. 48, 65–79. Semple, K.T., Doick, K.J., Jones, K.C., Burauel, P., Craven, A., Harms, H., 2004. Peer reviewed: defining bioavailability and bioaccessibility of contaminated soil and sediment is complicated. Environ. Sci. Technol. 38, 228A–231A. Serdar, C.M., Gibson, D.T., Munnecke, D.M., Lancaster, J.H., 1982. Plasmid involvement in parathion hydrolysis by Pseudomonas diminuta. Appl. Environ. Microbiol. 44, 246–249. Sethunathan, N., 1973. Microbial degradation of insecticides in flooded soil and in anaerobic cultures. Residue Rev. 47, 143–165. Sethunathan, N., Yoshida, T., 1973. A Flavobacterium sp. that degrades diazinon and parathion. Can. J. Microbiol. 19, 873–875. Sethunathan, N., Bautista, E.M., Yoshida, T., 1969. Degradation of benzene hexachloride by a soil bacterium. Can. J. Microbiol. 15, 1349–1354. Shafiani, S., Malik, A., 2003. Tolerance of pesticides and antibiotic resistance in bacteria isolated from wastewater-irrigated soil. World J. Microbiol. Biotechnol. 19, 897–901. Shaw, W.V., Leslie, A.G.W., 1991. Chloramphenicol acetyltransferase. Annu. Rev. Biophys. Chem. 20, 363–386. Sherchan, S.P., Bachoon, D.S., Otero, E., Ramsubhag, A., 2013. Molecular detection of atrazine catabolism gene atzA in coastal waters of Georgia, Puerto Rico and Trinidad. Mar. Pollut. Bull. 69, 215–218. Siddaramappa, R., Rajaram, K.P., Sethunathan, N., 1973. Degradation of parathion by bacteria isolated from flooded soil. Appl. Environ. Microbiol. 26, 846–849. Singh, B.K., 2009. Organophosphorus-degrading bacteria: ecology and industrial applications. Nat. Rev. Microbiol. 7, 156–164. Singh, B.K., Walker, A., 2006. Microbial degradation of organophosphorus compounds. FEMS Microbiol. Rev. 30, 428–471. Singh, B.K., Kuhad, R.C., Singh, A., Lal, R., Tripathi, K.K., 1999. Biochemical and molecular basis of pesticide degradation by microorganisms. Crit. Rev. Biotechnol. 19, 197–225. Somasundaram, L., Coats, J.R., Racke, K.D., Stahr, H.M., 1990. Application of the Microtox system to assess the toxicity of pesticides and their hydrolysis metabolites. Bull. Environ. Contam. Toxicol. 44, 254–259. Sommer, M.O., Church, G.M., Dantas, G., 2010. The human microbiome harbors a diverse reservoir of antibiotic resistance genes. Virulence 1, 299–303. Staley, Z.R., Senkbeil, J.K., Rohr, J.R., Harwood, V.J., 2012. Lack of direct effects of agrochemicals on zoonotic pathogens and fecal indicator bacteria. Appl. Environ. Microbiol. 78, 8146–8150. Staley, Z.R., Harwood, V.J., Rohr, J.R., 2015. A synthesis of the effects of pesticides on microbial persistence in aquatic ecosystems. Crit. Rev. Toxicol. 45, 813–836. Steffan, J.J., Brevik, E.C., Burgess, L.C., Cerda, A., 2018. The effect of soil on human health: an overview. Eur. J. Soil Sci. 69, 159–171. Sun, C., Shu, K., Wang, W., Ye, Z., Liu, T., Gao, Y., Zheng, H., He, G., Yin, Y., 2014. Encapsulation and controlled release of hydrophilic pesticide in shell cross-linked nanocapsules containing aqueous core. Int. J. Pharm. 463, 108–114. Sutherland, T.D., Horne, I., Harcourt, R.L., Russell, R.J., Oakeshott, J.G., 2002. Isolation and characterization of a Mycobacterium strain that metabolizes the insecticide endosulfan. J. Appl. Microbiol. 93, 380–389. Suwa, Y., Wright, A.D., Fukimori, F., Nummy, K.A., Hausinger, R.P., Holben, W.E., Forney, L.J., 1996. Characterization of a chromosomally encoded 2,4-dichlorophenoxyacetic acid/alpha-ketoglutarate dioxygenase from Burkholderia sp. strain RASC. Appl. Environ. Microbiol. 62, 2464–2469. Swiecilo, A., Zych-Wezyk, I., 2013. Bacterial stress response as an adaptation to life in a soil environment. Pol. J. Environ. Stud. 22, 1577–1587. Thal, L.A., Zervos, M.J., 1999. Occurrence and epidemiology of resistance to virginiamycin and streptogramins. J. Antimicrob. Chemother. 43, 171–176. Thomas, J.E., Ou, L.T., All-Agely, A., 2008. DDE remediation and degradation. Rev. Environ. Contam. Toxicol. 194, 55–69. Tomizawa, M., Casida, J.E., 2005. Neonicotinoid insecticide toxicology: mechanisms of selective action. Annu. Rev. Pharmacol. Toxicol. 45, 247–268. Tsvetkov, N., Samson-Robert, O., Sood, K., Patel, H.S., Malena, D.A., Gajiwala, P.H., Maciukiewicz, P., Fournier, V., Zayed, A., 2017. Chronic exposure to neonicotinoids reduces honey bee health near corn crops. Science 356, 1395–1397. Tu, C.M., 1992. Effect of some herbicides on activities of microorganisms and enzymes in soil. J. Environ. Sci. Health A B27, 695–709. Udikovic-Kolic, N., Wichmann, F., Broderick, N.A., Handelsman, J., 2014. Bloom of resident antibiotic-resistant bacteria in soil following manure fertilization. Proc. Natl. Acad. Sci. U. S. A. 111, 15202–15207. UN, 2017. Globally Harmonized System of Classification and Labelling of Chemicals (GHS). Seventh revised edition. United Nations, New York and Geneva https://

189

www.unece.org/fileadmin/DAM/trans/danger/publi/ghs/ghs_rev07/English/ST_SG_ AC10_30_Rev7e.pdf. UNEP, 2017. The 16 New POPs. An introduction to the chemicals added to the Stockholm Convention as Persistent Organic Pollutants by the Conference of the Parties. Stockholm Convention Secretariat, United Nations Environment, Geneva, Switzerlandwww.pops.int. US EPA, 2005. Reregistration eligibility document for methyl salicylate. OPP Chemical Code: 076601 United States Environmental Protection Agency, Washington D.C https://archive.epa.gov/pesticides/reregistration/web/pdf/methyl_salicylate_red.pdf. US EPA, 2014. Registration of Enlist Duo. United States Environmental Protection Agency https://www.epa.gov/ingredients-used-pesticide-products/registration-enlist-duo, Accessed date: 20 July 2018. US EPA, 2017. EPA data requirement for registration of antimicrobial pesticides: part 158W. Environmental Protection Agency, Office of Pesticide Program, Washington DChttps://www.epa.gov/pesticide-registration/epa-data-requirements-registrationantimicrobial-pesticides-part-158w. Velmurugan, G., Ramprasath, T., Swaminathan, K., Mithieux, G., Rajendhran, J., Dhivakar, M., Parthasarathy, A., Babu, D.D., Thumbuaj, L.J., Freddy, A.J., Dinakaran, V., Puhari, S.S., Rekha, B., Christy, Y.J., Anusha, S., Divya, G., Suganya, K., Meganathan, B., Kalyanaraman, N., Vasudevan, V., Kamaraj, R., Karthik, M., Jeyakumar, B., Abhishek, A., Paul, E., Pushpanathan, M., Rajmohan, R.K., Velayutham, K., Lyon, A.R., Ramasamy, S., 2017. Gut microbial degradation of organophosphate insecticidesinduces glucose intolerance via gluconeogenesis. Genome Biol. 18, 8. Venkateswarlu, K., Sethunathan, N., 1978. Degradation of carbofuran in rice soils as influenced by repeated applications and aerobic conditions following anaerobiosis. J. Agric. Food Chem. 26, 1148–1151. Venkateswarlu, K., Sethunathan, N., 1979. Metabolism of carbofuran in rice strawamended and unamended rice soils. J. Environ. Qual. 8, 365–368. Vieweger, A., Doring, T.F., 2015. Assessing health in agriculture-towards a common research framework for soils, plants, animals, humans and ecosystems. J. Sci. Food Agric. 95, 438–446. Vuilleumier, S., 1997. Bacterial glutathione S-transferases: what are they good for? J. Bacteriol. 179, 1431–1441. Walsh, F., Amyes, S.G.B., Duffy, B., 2013. Challenging the concept of bacteria subsisting on antibiotics. Int. J. Antimicrob. Agents 41, 558–563. Wang, C.J., Liu, Z.Q., 2007. Foliar uptake of pesticides-present status and future challenge. Pestic. Biochem. Physiol. 87, 1–8. Watts, J.E.M., Schreier, H.J., Lanska, L., Hale, M.S., 2017. The rising tide of antimicrobial resistance in aquaculture: sources, sinks and solutions. Mar. Drugs 15 (6), 158. Weinstock, G.M., 2001. Microbial genetics. In: Brenner, S., Miller, J.H. (Eds.), Encylopedia of Genetics. Academic Press. ISBN: 978-0-12-227080-2, pp. 1191–1196. WHO, 2010. Guidelines on public health pesticide management policy. World Health Organization, Regional Office for South-East Asia, WHO Pesticide Evaluation Scheme (WHOPES). WHO, Geneva http://www.who.int/whopes/resources/SEA_CD_214.pdf. WHO, 2016. The public health impact of chemicals: knowns and unknowns. International Programme on Chemical Safety. World Health Organization, Geneva, Switzerlandhttp://apps.who.int/iris/bitstream/handle/10665/206553/WHO_FWC_ PHE_EPE_16.01_eng.pdf. WHO, 2017. Global Priority List of Antibiotic-Resistant Bacteria to Guide Research, Discovery, and Development of New Antibiotics. Publication date: 27 February 2017. World Health Organization, Geneva, Switzerland, p. 7. www.who.int/ipcs/en/ http://www. who.int/medicines/publications/global-priority-list-antibiotic-resistant-bacteria/en/. Woappi, Y., Gabani, P., Singh, A., Singh, O.V., 2016. Antibiotrophs: the complexity of antibiotic-subsisting and antibiotic-resistant microorganisms. Crit. Rev. Microbiol. 42, 17–30. Wright, G.D., 2007. The antibiotic resistome: the nexus of chemical and genetic diversity. Nat. Rev. Microbiol. 5, 175–186. Wright, A.C., Danyluk, M.D., Otwell, W.S., 2009. Pathogens in raw foods: what the salad bar can learn from the raw bar. Curr. Opin. Biotechnol. 20, 172–177. Wuichet, K., Zhulin, I.B., 2010. Origins and diversification of a complex signal transduction system in prokaryotes. Sci. Signal. 3, ra50. Xia, X., Sun, B., Gurr, G.M., Vasseur, L., Xue, M., You, M., 2018. Gut microbiota mediate insecticide resistance in the Diamonback moth, Plutella xylostella (L.). Front. Microbiol. 9, 25. Ximenes, E., Hoagland, L., Ku, S., Li, X., Ladisch, M., 2017. Human pathogens in plant biofilms: formation, physiology, and detection. Biotechnol. Bioeng. 114, 1403–1418. Yan, J., Guan, H., Yu, J., Chi, D., 2013. Acetylcholinesterase biosensor based on assembly of multiwall carbon nanotubes onto liposome bioreactors for detection of organophosphates pesticides. Pestic. Biochem. Physiol. 105, 197–202. Yang, H., Carr, P.D., McLoughlin, S.Y., Liu, J.W., Horne, I., Qiu, X., Jeffries, C.M.J., Russell, R.J., Oakeshott, J.G., Ollis, D.L., 2003. Evolution of an organophosphate-degrading enzyme: a comparison of natural and directed evolution. Protein Eng. Des. Sel. 16, 135–145. Zacharia, J.T., 2011. Identity, physical and chemical properties of pesticides. In: Stoytcheva, M. (Ed.), Pesticides in the Modern World – Trends in Pesticides Analysis. InTech Europe, Rijeka, Croatia. ISBN: 978-953-307-437-5. Zhang, M., Liang, Y., Song, A., Yu, B., Zeng, X., Chen, M.-S., Yin, H., Zhang, X., Sun, B., Fan, F., 2017. Loss of soil microbial diversity may increase insecticide uptake by crop. Agric. Ecosyst. Environ. 240, 84–91.