Biodegradation of the Neonicotinoid Insecticide Acetamiprid by

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Agricultural and Environmental Chemistry

Biodegradation of the neonicotinoid insecticide acetamiprid by actinomycetes Streptomyces canus CGMCC 13662 and characterization of the novel nitrile hydratase involved Ling Guo, Wenwan Fang, Leilei Guo, Chuanfei Yao, Yunxiu Zhao, Feng Ge, and Yijun Dai J. Agric. Food Chem., Just Accepted Manuscript • DOI: 10.1021/acs.jafc.8b06513 • Publication Date (Web): 08 May 2019 Downloaded from http://pubs.acs.org on May 8, 2019

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Journal of Agricultural and Food Chemistry

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Biodegradation of the neonicotinoid insecticide acetamiprid

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by actinomycetes Streptomyces canus CGMCC 13662 and

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characterization of the novel nitrile hydratase involved

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Ling Guoa · Wen-Wan Fanga · Lei-Lei Guoa · Chuan-Fei Yaoa · Yun-Xiu

5

Zhaoa · Feng Geb* · Yi-Jun Daia*

a

Jiangsu Key Laboratory for Microbes and Functional Genomics, Jiangsu Engineering and

Technology Research Center for Industrialization of Microbial Resources, College of Life Science, Nanjing Normal University, Nanjing 210023, People’s Republic of China b

Nanjing Institute of Environmental Sciences, Ministry of Environmental Protection,

Nanjing 210042, People’s Republic of China W. W. Fang is the co-first author 1 / 37 ACS Paragon Plus Environment


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ABSTRACT: Neonicotinoid insecticide pollution in soil and water poses serious

2

environmental risks. Microbial biodegradation is an important neonicotinoid insecticide

3

degradation pathway in the environment. In this study, 70.0% of the acetamiprid in a 200

4

mg/L solution was degraded by actinomycetes Streptomyces canus CGMCC 13662

5

(isolated from soil) in 48 h, and the acetamiprid degradation half-life was 27.7 h.

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Acetamiprid was degraded to IM-1-2 ((E)-1-(1-(((6-chloropyridin-3-yl)methyl)(methyl)

7

amino)ethylidene)urea) through hydrolysis of the cyanoimine moiety. Gene cloning and

8

over-expression indicated that a novel nitrile hydratase with three unusual subunits (AnhD,

9

AnhE, and AnhA) without accessory protein mediated IM-1-2 formation. The purified

10

nitrile hydratase responsible for degrading acetamiprid had a Km of 5.85 mmol/L and a Vmax

11

of 15.99 U/mg. A homology model suggested that AnhD-Glu56 and AnhE-His21 play

12

important roles in the catalytic efficiency of the nitrile hydratase. S. canus CGMCC 13662

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could be used to remediate environments contaminated with acetamiprid.

14 15

KEYWORDS: acetamiprid; biodegradation; nitrile hydratase; Streptomyces canus

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Introduction

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Neonicotinoid insecticides (e.g., acetamiprid (ACE), clothianidin, dinotefuran, imidacloprid,

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nitenpyram, thiacloprid (THI), and thiamethoxam) are neuro-active insecticides that are

4

used extensively (added to soil or applied as seed dressings or sprays) to control important

5

agricultural crop pests.1,2 Imidacloprid was made commercially available in the early 1990s,

6

and neonicotinoid insecticide use has increased considerably since then. Neonicotinoid

7

insecticides are now sold in larger amounts than any other insecticides around the world.1

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However, neonicotinoid insecticides have recently been linked to adverse ecological effects,

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including honeybee colony collapse disorder, threats to aquatic invertebrate species

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survival, and widespread decreases in butterfly and insectivorous bird numbers.2-5 Most

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neonicotinoid insecticides are water-soluble and break down slowly in the environment,

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and >80% of neonicotinoid insecticide residues in treated fields eventually enters surface or

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ground water, contaminating the aquatic environment and harming aquatic arthropods.6-8

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The structure of ACE (N-[(6-chloro-3-pyridyl)methyl]-Nʹ-cyano-N-methylacetamidine)

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is shown in Figure S1. ACE is an insecticide in the chloronicotinyl subclass of the

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neonicotinoid pesticide class, and contains a cyanoguanidine moiety, which is a

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pharmacophore.9,10 ACE is widely used to protect crops because it can be used to control a

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broad range of insects but is only moderately toxic to honey bees.11 However, it has

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recently been found that ACE induces clinical symptoms (e.g., decreased body weight,

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respiratory depression, and hepatic effects) in mice.6,12 Long-term exposure to ACE posed

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continual health hazards to Oreochromis mossambicus, implying that ACE will pose risks

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to humans consuming fish that have been exposed to ACE.13 In a recent study it was

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proposed that ACE may decrease human fertility.14 ACE has a solubility of 4.25 g/L in 3 / 37 ACS Paragon Plus Environment


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water at 25 °C, so is mobile in soil and can permeate ground and surface water. ACE in the

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environment therefore poses risks to sensitive aquatic invertebrates and whole

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ecosystems.3,15 Much attention has therefore been paid to ACE residues in the environment

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and methods for assessing ACE degradation.

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Photocatalytic degradation (using TiO2 and light produced by a Xe lamp) and oxidation

6

during Fenton reactions have been used to remove ACE from water samples.16,17 Physical

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and chemical degradation methods required extreme conditions, are expensive, and cause

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environmental contamination, but microbial biodegradation is environmental friendly and

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economical. Many microbes have been found to degrade ACE. For example, the bacterium

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Stenotrophomonas maltophilia CGMCC 1.1788 degrades ACE through N-demethylation to

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give IM-2-1 (N-((6-chloropyridin-3-yl)methyl)acetamide).18 The yeast Rhodotorula mucila

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ginosa IM-2 degrades ACE through oxidative cleavage of the cyanoguanidine moiety to

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give IM-1-3 (N-((6-chloropyridin-3-yl)methyl)-N-methylacetamide).19 The bacteria Vario

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vorax boronicumulans CGMCC4969 and Ensifer meliloti CGMCC 7333 degrade ACE

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through hydration of the cyano group to give the N-carbamoylimine derivative IM-1-2 ((E)

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-1-(1-(((6-chloropyridin-3-yl)methyl)(methyl)amino)ethylidene)urea).20,21Pigmentiphaga sp.

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AAP-122 and Pigmentiphaga sp. D-2, 23 Rhodococcus sp. BCH2, 24 and Stenotroph omonas

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sp. THZ-XP25 degrade ACE to give IM-1-4 ((6-chloropyrid in-3-yl)-N-methylmethana

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mine) (Figure S1).

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It has been found in studies focused on enzymes that the cobalt-type nitrile hydratase

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(NHase) EC 4.2.1.84 is responsible for converting ACE into IM-1-2.20,21 Several NHases

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with different subunit organizations are shown in Figure 1. The ACE-converting NHases in

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V. boronicumulans CGMCC4969 and E. meliloti CGMCC 7333 belong to type “a”, in

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which the encoding gene order is anhA (α-subunit) followed by anhB (β-subunit) and anhC 4 / 37 ACS Paragon Plus Environment

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(activator protein) and in which AnhC is essential to the NHase maturity and catalytic

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activity.26 The acetonitrile-catabolic NHase (ANHase) in Rhodococcus jostii RHA1 belongs

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to type “b”, in which the AnhC acts as a metallochaperone and is between the AnhA and

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AnhB.27,28 The L-NHases in Rhodococcus rhodochrous J1 are type “c” NHases, and have

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the gene structure order anhB, anhA, anhC, with AnhC encoded by anhC gene assisting

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with metal insertion and oxidation of the two equatorial cysteine residue active sites.28,29 A

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type “d” NHase with a single subunit fusing AnhB and AnhA with no AnhC present is

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found in the eukaryote Monosiga brevicollis, which has a stronger affinity for aromatic than

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aliphatic nitrile substrates.30,31

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Many microbes have been found to exhibit NHase activities, and their roles in industrial

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syntheses of amides are well documented. However, the catalytic efficiencies and

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biochemical and structural properties of NHases related to the biodegradation of

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nitrile-containing organic contaminators (particularly nitrile-containing pesticides used

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around the world) have been studied little. In this study, we found that the actinomycetes

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Streptomyces canus CGMCC 13662, isolated from soil, can rapidly degrade ACE, and we

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identified a novel NHase that degrades ACE to give IM-1-2. We investigated the functions

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of the NHase subunits and assessed the possible NHase maturation mechanism. The results

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will provide a theoretical basis for remediating environmental media contaminated with

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nitrile-containing pesticides.

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MATERIALS AND METHODS

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Chemicals and Media. The ACE and THI used were both >98% purity, and were

22

provided by Dr. Haijun Ma of the Jiangsu Pesticide Research Institute, Nanjing, China. A

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standard substance of IM-1-2 was prepared using a method described by Zhou et al.21



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IM-1-4 (purity >95%) was purchased from Sigma Aldrich (Shanghai, China). Acetonitrile,

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used for high-performance (HP) liquid chromatography (LC), was of HPLC grade and was

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purchased from Merck (Darmstadt, Germany). All other reagents were of analytical grade

4

and were purchased from Sangon Biotech (Shanghai, China).

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The enrichment medium was a mineral salt medium (pH 7.5), a solution of 10.0 g

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glucose, 1.36 g KH2PO4, 2.13 g Na2HPO4, and 0.50 g MgSO4·7H2O in 1.0 L deionized

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water. The actinomycetes culture medium was ISP4 medium (pH 7.2), a solution of 10.0 g

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soluble starch, 1.0 g K2HPO4, 1.0 g MgSO4.7H2O, 1.0 g NaCl, 2.0 g (NH4)2SO4, 2.0 g

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CaCO3, 0.001 g FeSO4.7H2O, and 0.001 g MnCl2.7H2O in 1.0 L deionized water. If

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necessary, 0.1 mmol/L CoCl2 was added to the ISP4 medium to allow active NHase to form.

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Gauze’s synthetic medium (pH 7.2) was prepared by dissolving 20.0 g soluble starch, 1.0 g

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KNO3, 0.50 g K2HPO4, 0.50 g MgSO4.7H2O, 0.50 g NaCl, and 0.01 g FeSO4.7H2O in 1.0 L

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deionized water. Lysogeny broth medium (LB; at pH 7.2), prepared by dissolving 10.0 g

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peptone, 5.0 g yeast extract, and 10.0 g NaCl in 1.0 L deionized water, was used for

15

isolating, purifying, and incubating bacteria.

16 17

Isolating Microbes from Soil and Identifying the Microbes. The detailed method of isolating microbes from soil was shown in the Supporting Information.

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The 16S rRNA gene was amplified by performing the colony polymerase chain reaction

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and then sequenced by Sangon Biotech. The primers (K1 and K2) used for the 16S rRNA

20

gene amplification and sequencing processes are listed in Table S1 in the Supporting

21

Information.

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Biodegradation of ACE by Resting Cells and Transformation in

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Growing Cultures. The degradation of ACE by resting actinomycetes cells was 6 / 37 ACS Paragon Plus Environment

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assessed by streak inoculating the isolate onto a plate covered with agar containing Gauze’s

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medium and then incubating the plate at 30 °C for about 7 d until spores appeared. The

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spores were then added to a 250 mL flask containing 50 mL ISP4 broth containing 0.1

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mmol/L CoCl2. The broth was incubated under the conditions described above for 4 d, and

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then the bacterial cells were harvested. A 0.30 g (wet weight) aliquot of the cells was

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re-suspended in 5 mL of sodium phosphate buffer containing 200 mg/L ACE. Control tests

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were performed using broth not containing cells or ACE. The biodegradation of ACE by

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growing actinomycetes cultures was assessed by adding an aliquot of the spore suspension

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to 50 mL ISP4 broth containing 0.1 mmol/L CoCl2 and 200 mg/L ACE in a 250 mL flask.

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The other methods used to assess ACE degradation by resting and growing actinomycetes

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cells were described in a previous publication.21

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Biodegradation of ACE in Soil by the Isolated Microbes. In order to assess

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remediation efficiency of the isolated microbes for ACE residues in soil, we explored the

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biodegradation dynamics of ACE in soil by the isolated microbes. Soil samples were

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collected from the Xianlin campus of Nanjing Normal University, China. Each sample was

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dried in air, ground, passed through a sieve with 2 mm pores, and then dry-heat-sterilized

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for 2 h. Degradation of ACE in the soil by actinomycetes was assessed following a

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previously published method.32 The isolated microbes were cultivated using the method

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described above for one week, then the mycelia were collected by vacuum filtration and 0.2

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g wet mycelia were suspended in 2 mL 0.2 mol/L sodium phosphate buffer (pH 7.5). The

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suspension was added evenly to 10 g of soil. The final ACE content of the soil was 5 mg/kg

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soil. The moisture content of the soil was 30%. Recovery limit of detection (LOD) and

23

limit of quantification (LOQ) were considered as the method of standardization and



1

validation of ACE residue analysis in soil, of which the definition and method were

2

following the method published by Lee et al.33

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HPLC and LC Mass Spectrometry (MS). An Agilent 1200 HPLC system

4

equipped with an Agilent G1314A UV detector (Agilent Technologies, Santa Clara, CA,

5

USA) was applied for analysis of ACE and its metabolite. An Agilent 1290 infinity LC

6

with a G1315B diode array detector and an Agilent 6460 triple-quadrupole LC-MS system

7

equipped with an electrospray ion source (Agilent Technologies) was used for LC-MS

8

analysis. The column was an HC-C18 column (4.6 × 250 mm, 5 μm particle size) equipped

9

with a reverse-phase HC-C18 guard column (4.6×12.5 mm, Agilent Technologies). The

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HPLC system flow rate was 1 mL/min, and the mobile phase was a 70:30 mixture of water

11

(containing 0.01% acetic acid) and acetonitrile. The same mobile phase was used for the

12

LC-MS analyses, but the flow rate was 0.6 mL/min. The diode array detector was used at a

13

wavelength of 235 nm.

14

Cloning the NHase Gene from the Strain 2-18 and Construction of the

15

Plasmid. Total genomic DNA was extracted from the isolate using a TaKaRa MiniBEST

16

Bacteria Genomic DNA Extraction Kit Ver. 3.0 (TaKaRa, Dalian, China). High-fidelity

17

PrimeSTAR HS DNA Polymerase (TaKaRa catalog no. DR010A) was used. The primers

18

used to clone the NHase are shown in Table S1 of Supporting Information.

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Plasmids Q0, Q1, Q2, Q3, and Q4 were constructed in the base of vector pET-28a(+) to

20

allow the function of each NHase subunit to be studied. The primers and cloning strategies

21

used to construct the plasmids are shown in Table S2 and Figure S2 in the Supporting

22

Information. The NHase gene recombinant pET-28a was constructed following the

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ClonExpress II one step cloning kit (Vazyme Biotech, Nanjing, China) protocol. The 8 / 37 ACS Paragon Plus Environment

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recombinant pET-28a was verified by DNA sequencing (Sipkin Biotech, Nanjing, China).

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Expression and Purification of Recombinant NHase in Escherichia coli

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Rosetta. The E. coli Rosetta competent cells and calcium chloride transformed cells were

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prepared following a method published by Sun et al.20 NHase overexpression in E. coli

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Rosetta was analyzed and NHase was purified by His-tag affinity chromatography

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following methods described by Rzeznicka et al.34 Sodium dodecyl sulphate

7

polyacrylamide gel electrophoresis (SDS-PAGE) was performed to assess protein

8

expression. Gels were stained with Coomassie Brilliant Blue to allow proteins to be

9

detected.

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The bands corresponding to the NHase subunits were identified by Western blot analysis

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and matrix-assisted laser desorption ionization time-of-flight mass spectrometry

12

(MALDI-TOF MS), performed by Sangon Biotech. Overexpressed NHase characterization

13

and Western blot analyses were performed following previously described methods.20,26

14

The solubility of the overexpressed NHase was determined using a previously described

15

method.20 One unit (U) of NHase activity was deﬁned as the amount of enzyme that

16

catalyzed the formation of 1 μmol IM-1-2 in 1 min.

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Homology Model for S. canus CGMCC 13662 NHase. A homology model

18

for S. canus CGMCC 13662 NHase (ScNHase) was developed based on the crystal

19

structure of Co-type nitrile hydratase from Pseudomonas putida (PDB code: 3QXE). The

20

SWISS-MODEL workspace was used to build and evaluate the ScNHase homology model,

21

and GMQE35 and QMEAN36 were used to assess the qualities of the constructed NHase

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structure. Alignments of the structures between the ScNHase subunit and templates (3QXE

23

A and B chains) were implemented using the PyMOL program. 9 / 37 ACS Paragon Plus Environment


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Characteristics of the NHase. The optimal temperature for the degradation of

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ACE by NHase was determined by performing tests in 0.2 mol/L phosphate-buffered saline

3

at pH 7.0 containing 500 mg/L ACE and 0.1 mmol/L CoCl2 at temperatures between 20

4

and 60 °C. The optimal pH for the degradation of ACE by NHase was determined by

5

performing tests at the optimal temperature. Aliquots of phosphate-buffered saline at

6

different pH values between pH 4 and 9 and containing 500 mg/L ACE were used.

7

The effects of the pH on NHase stability were assessed by performing preliminary tests

8

in which mixtures of NHase and phosphate-buffered saline at between pH 4 and 9 were

9

incubated at 30 C for 2 h with shaking at 220 rpm. The effects of the temperature on

10

NHase stability were assessed by performing preliminary tests in which mixtures of NHase

11

and phosphate-buffered saline at pH 7 were incubated at between 20 and 60 °C. The

12

influences of various metal ions on ACE degradation by NHase were investigated using

13

standard reaction mixtures containing 0.1 mmol/L CaCl2, CuCl2, FeCl3, MnCl2, MoCl5, or

14

ZnCl2. Each test of NHase activity was performed at the optimal temperature and pH.

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The activities of NHase toward 13 substrates were determined. The substrates were ACE,

16

azoxystrobin,

benzonitrile

(BN),

17

cyhalofop-butyl,

18

(R)-(+)-4-methylmandelo nitrile, and THI. Each test was performed using a 1 mL solution

19

containing NHase and 50 mg/L of a substrate. The HPLC conditions used have been

20

published previously.26 Kinetics parameters for the reaction between NHase and ACE were

21

determined using the Eadie-Hofstee graphical method using the linear relationship between

22

the enzyme reaction velocity (V) and the enzyme reaction velocity to substrate

23

concentration ratio (V/[S]).

dichlobenil,

bromoxynil, fipronil,

chlorfenapyr,

fludioxonil,


2-cyanopyridine,

indole-3-acetonitrile,

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RESULTS AND DISCUSSION

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Isolation and Identification of ACE-degrading Microbes. About 120

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colonies growing on the mineral salt medium agar plate were streaked onto LB agar plates

4

and their ACE-degradation abilities tested by HPLC. Three strains of bacteria had low

5

ACE-degradation abilities, and a strain named gl2-18 had a comparatively high

6

ACE-degradation ability. Morphological observations indicated that gl2-18 had a colonial

7

morphology typical of actinomycetes cultivated in Gauze’s agar medium. Actinomycetes

8

have rarely been found to degrade neonicotinoid insecticides.24 The related enzyme

9

research of Rhodococcus sp. BCH2 biodegrading ACE have not been reported,24 so ACE

10

degradation by strain gl2-18 was studied further. Nucleotide blast analysis and phylogenic

11

tree analysis of the 16S rRNA gene sequences indicated that gl2-18 clustered with S. canus

12

strain S02 (HQ850405) (Figure 2A). Strain gl2-18 was deposited in the China General

13

Microbiological Culture Collection Center (CGMCC) (Beijing, China) under the accession

14

number CGMCC 13662. The partial 16S rRNA gene sequence of S. canus CGMCC 13662

15

was deposited in the Genbank database under accession number MG657358.

16

Identification of Metabolites and the Kinetics of ACE Degradation by S.

17

canus CGMCC 13662. Resting S. canus CGMCC 13662 cells degraded ACE and

18

caused a polar metabolite with a HPLC retention time of 3.7 min to form (Figure 3C), but

19

the metabolite did not form in the control samples containing only ACE or only bacteria

20

(Figure 3A and 3D). The LC-MS spectrum for metabolite P1 (Figure 3C) had a protonated

21

ion (M+H) at m/z 241, a M−NH2 fragment ion at m/z 224, an unknown fragment ion at m/z

22

198, and a M−C4H8N3O fragment ion at m/z 126. ACE had a protonated ion (M+H) at m/z



1

223 (Figure 3A). P1 had the same retention time and mass as the N-carbamoylimine

2

derivative of ACE called IM-1-2. ACE degradation by the resting S. canus CGMCC 13662

3

cells therefore involved the hydration of ACE to give IM-1-2.

4

The kinetics curve for the degradation of ACE by resting S. canus CGMCC 13662 cells

5

(Figure 3F) indicated that the actinomycetes degraded 0.79 mmol/L ACE (degradation rate

6

87.6%) in 4 d and produced 0.65 mmol/L IM-1-2. The molar ACE to IM-1-2 conversion

7

rate was 82.6%, indicating that hydrolysis of ACE to give IM-1-2 was the main metabolic

8

pathway. The ACE degradation half-life was 27.7 h, which was shorter than the half-life of

9

62.4 h for the degradation of ACE by E. meliloti CGMCC 7333.21

10

The degradation of ACE by growing S. canus CGMCC 13662 cells was also studied. As

11

shown in Figure 3E, the spores grew to form a mycelium pellet in the first 2 d and

12

decreased the ACE concentration from 0.99 to 0.83 mmol/L. ACE was then degraded

13

rapidly, the concentration decreasing to 0.27 mmol/L after 4 d and 0.09 mmol/L after 6 d.

14

The ACE concentration decreased by 0.73 mmol/L in 4 d, but only 0.12 mmol/L of IM-1-2

15

was formed (Figure 3E). The only product of the degradation of ACE by resting cells was

16

IM-1-2, but the growing cells also produced product P2, which had a retention time of 2.70

17

min. In the LC-MS spectrum of P2 (Figure 3B), the parent ion was at m/z 157 and there

18

was an acetonitrile adduct at m/z 198, a fragment ion at m/z 126 (the same m/z ratio as for

19

IM-1-4, see Figure 3B), and an unknown fragment ion at m/z 105. P2 had the same

20

retention time and mass as IM-1-4, so we concluded that P2 was IM-1-4. The IM-1-4

21

concentrations on days 4 and 6 were 0.51 and 0.65 mmol/L, respectively. It has previously

22

been found that IM-1-4 is spontaneously produced through the hydrolysis of IM-1-2 and

23

that the conversion of IM-1-2 into IM-1-4 is accelerated under acidic conditions (data not

24

shown). 21 The ISP4 medium inoculated with actinomycetes started at pH 7.2 but decreased 12 / 37 ACS Paragon Plus Environment

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Biodegradation of the Neonicotinoid Insecticide Acetamiprid by

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