Status Report on Neonicotinoid Pollution in Japan: Emphasis on Human, Animal, and Environmental Health

ABSTRACT

Neonicotinoid insecticides are exclusively used for insect control in many parts of the globe. Since early 1992, various neonicotinoid active compounds have been used for pest control in Japan. At present, all seven major neonicotinoid active compounds, imidacloprid, acetamiprid, dinotefuran, clothianidin, thiamethoxam, thiacloprid, and nitenpyram, are used in Japan, and the total shipment amount of these chemicals in the country was estimated 337.9 tons in 2022. Due to the massive neonicotinoid application in Japan, concerns have been raised about their potential infiltrations into nontarget ecosystems and their effects on human and ecological species. These concerns have inspired several studies, extensively investigating the environmental contamination levels and exposure tendencies in humans and several other invertebrate species. The current status report harmonized the up-to-date data on neonicotinoids in Japan and identified the incoherencies or knowledge gaps existing in the neonicotinoid literature reported in Japan. The report further summarized extensive neonicotinoid publications globally to illustrate the distributions of various neonicotinoid compounds in agricultural soils and surface waters and their exposures in humans and nontarget animal species. This report represents a few perspectives with detailed data from Japan into perspective.

INTRODUCTION

Over the years, various synthetic pesticides have been employed in several pest management programs worldwide, with society seemingly accepting their costs-benefit balance. Although these pesticides offer several benefits, their potential deleterious impacts on ecosystem biodiversity have raised serious concerns globally. Especially, legacy pesticides, including organochlorines and organophosphates, are well known to elicit high toxicological manifestations in humans and other nontarget vertebrates (Karami-Mohajeri and Abdollahi, 2011). Organochlorine pesticides, in particular, were red-listed and banned because of their high environmental persistence and bioaccumulation and biomagnification tendencies in mammalian systems (Mrema et al., 2013). Besides, many legacy pesticides have been found to cause resistance in the species they are intended to treat (Metcalf, 1989; Ndiath, 2019). As a result of pest resistance, high cumulative exposure trends of organophosphorus and carbamates, and as a replacement for cholinesterase inhibitors with strong negative impacts on children’s neural development (Eskenazi et al., 1999; Morrissey et al., 2015; Craddock et al., 2019), neonicotinoid insecticides were developed and commercialized in the early 1990s.

Neonicotinoid insecticides currently reign as the most preferred insecticides globally (Marfo et al., 2015; Anai et al., 2021). Several neonicotinoid active compounds are frequently used to control biting and chewing insects in homes, veterinary science, and agriculture (Craddock et al., 2019). As systemic insecticides, neonicotinoid residues are perceived to have percolated into various compartments of the environment. For the past 2 decades, a large suite of studies has focused on neonicotinoid infiltrations into aquatic ecology, terrestrial environments, food, and biota (MacDonald et al., 2018; Takeuchi and Nishizawa, 2018; Nishino et al., 2018; Yamamuro et al., 2019; Graves et al., 2023). Moreover, a couple of biomonitoring studies have consistently reported neonicotinoids in humans and other nontarget vertebrates (Ueyama et al., 2015; Harada et al., 2016; Kabata et al., 2016; Osaka et al., 2016; Honda et al., 2019; Ikenaka et al., 2019; Ospina et al., 2019; Tao et al., 2019a; Tao et al., 2019b; Zhang et al., 2019b; Li and Kannan, 2020; Li et al., 2020c; Song et al., 2020; Wang, et al., 2020a; Nimako et al., 2021; Oya et al., 2021; Taira et al., 2021; Nimako et al., 2022; Shinya et al., 2022). These findings indicate that neonicotinoid residues are everywhere within the environment and hence may inadvertently affect environmental quality, as well as human and animal health.

In Japan, extensive work has been done on environmental monitoring, exposure assessment, and risk characterization of neonicotinoid compounds in various environmental media. In addition, extensive biomonitoring studies have been carried out among various age groups in the Japanese population (Ueyama et al., 2015; Osaka et al., 2016; Harada et al., 2016; Ikenaka et al., 2019; Li and Kannan, 2020; Oya et al., 2021). The current status report harmonizes up-to-date data on neonicotinoids in Japan and identifies the incoherencies and knowledge gaps existing in the neonicotinoid literature reported in Japan; it encapsulates extensive global publications that show the distribution of neonicotinoid insecticides in agricultural soil and surface waters, as well as their exposure in humans and nontarget animal species. This report represents a few perspectives describing the detailed data from Japan.

MATERIALS AND METHODS

A literature evaluation was conducted to reveal the status of neonicotinoid contamination in Japanese ecosystems. Web sources such as PubMed, Google Scholar, Web of Science, etc., were explored for scientific literature on a wide range of topics on neonicotinoid insecticides, including biomonitoring, environmental monitoring, animal and human health impact assessment, and neonicotinoid invasion reports in the environment and food chain.

The literature search was completed using keywords including “pesticides,” “neonicotinoid insecticides,” “Japan,” “consumption,” “soil,” “food,” “water,” “human,” “wildlife,” “birds,” “contamination,” and “exposures.” The search language was either set to English or Japanese, but the publication date was not limited. The papers and reports obtained were scrutinized to determine their relevance to the overall structure of the current status report. Subsequently, the authors mined data from the selected publications. Figs. 2, 3, 4 and Tables 1 and 3, 4, 5 summarize some of the papers utilized.

Fig. 1 Chemical structures of neonicotinoids insecticides (Wang et al., 2017; Siviter and Muth, 2020)

Fig. 2A Total amount of neonicotinoids (tons) sold in Japan from 2016–2020. The pewter grey charts represent shipments in 2016, the stone grey charts represent shipments in 2018, and the black charts represent shipments in 2020. Data were deduced from the “Webkis-Plus: Database of chemicals” (https://www.nies.go.jp/kisplus/). The data for each chemical represents the summation of the amount of neonicotinoid compounds shipped to 47 Prefectures of Japan in a particular year

Fig. 2B Average amount of neonicotinoids (tons) shipped to the various Prefectures of Japan from 2016–2020. Data were deduced from the “Webkis-Plus: Database of chemicals” (https://www.nies.go.jp/kisplus/).

Fig. 3A Average urinary detection frequencies of neonicotinoids reported in 10 Japanese biomonitoring studies (n=2,102). Harada et al., 2016 (n=373), Anai et al., 2021 (n=109), Ichikawa et al., 2019 (n=65), Ikenaka et al., 2019 (n=46), Marfo et al., 2015 (n=85); Osaka et al., 2016 (n=223), Oya et al., 2021 (n=1,036); Ueyama et al., 2014 (n=52), Ueyama et al.,2019 (n=50), Nimako et al., 2022 (n=63)

Fig. 3B Average detection frequencies of neonicotinoids estimated for various age groups in 9 Japanese biomonitoring studies (n=2,039). The black charts represent neonicotinoid detection rates in infants less than 1.5 years; the burgandy chats represent neonicotinoid detection rates in children aged from 3 to 6 Years; and the grey charts represent neonicotinoid detection rates recoeded in Japanese adults. Harada et al., 2016 (n=373), Anai et al., 2021 (n=109), Ichikawa et al., 2019 (n=65), Ikenaka et al., 2019 (n=46), Marfo et al., 2015 (n=85); Osaka et al., 2016 (n=223), Oya et al., 2021 (n=1,036); Ueyama et al., 2014 (n=52), Ueyama et al., 2019 (n=50)

Fig. 4A Median urinary concentrations of neonicotinoids reported in Japanese subjects from 9 Japanese biomonitoring studies (n=2,037). A; Harada et al., 2016 (n=373), B; Anai et al., 2021 (n=109), C; Ikenaka et al., 2019 (n=46), D; Marfo et al., 2015 (n=85); E; Osaka et al., 2016 (n=223), F; Oya et al., 2021 (n=1,036); G; Ueyama et al., 2014 (n=52), H; Ueyama et al., 2019 (n=50), I; Nimako et al., 2022 (n=63)

Fig. 4B Maximum urinary concentrations of neonicotinoids reported in Japanese subjects from 9 Japanese biomonitoring studies (n=2,037). A; Harada et al., 2016 (n=373), B; Anai et al., 2021 (n=109), C; Ikenaka et al., 2019 (n=46), D; Marfo et al., 2015 (n=85); E; Osaka et al., 2016 (n=223), F; Oya et al., 2021 (n=1,036); G; Ueyama et al., 2014 (n=52), H; Ueyama et al., 2019 (n=50), I; Nimako et al., 2022 (n=63)

Table 1 Detection frequencies (%) and residual concentrations of neonicotinoids (ng/L) reported in river water bodies across countries

Country	Description	N	ACE	CLO	IMI	NIT	THI	THXM	DIN	dn-IMI	IMI-urea	dm-ACE	Reference
Detection Frequencies (%)
Japan	River waters	210.0	2.4	53.3	67.1	—	1.0	22.4	30.5	—	—	—	Sato et al., 2016
Japan	Tap water	21.0	—	38.1	47.6	—	—	—	9.5	—	—	—	Sato et al., 2016
Canada	Tile drains	50.0	8.0	88.0	12.0	0.0	2.0	58.0	2.0	—	—	—	Schaafsma et al., 2019
Canada	Ditches	119.0	5.0	95.0	13.0	0.0	2.0	50.0	3.0	—	—	—	Schaafsma et al., 2019
China	Yangtze river	120.0	100.0	64.0	100.0	73.0	87.0	95.0	—	—	—	—	Mahai et al., 2019
Philipines	Surface water	6.0	0.0	0.0	50.0	—	—	17.0	—	—	—	—	Bonmatin et al., 2021
Philipines	Surface water	2.0	0.0	0.0	50.0	—	—	50.0	—	—	—	—	Bonmatin et al., 2021
USA	River and streams	6.0	34.0	81.0	91.0	—	—	71.0	—	—	—	—	Berens et al., 2021
USA	Lakes	5.0	6.0	34.0	71.0	—	—	15.0	—	—	—	—	Berens et al., 2021
USA	Wells	7.0	48.0	48.0	45.0	—	—	53.0	—	—	—	—	Berens et al., 2021
Reported median neonicotinoid concentrations (ng/L)
Japan	River systems in Gifu	672.0	2.8	5.6	4.0	<LOQ	<LOQ	3.6	11.3	—	—	—	Hayashi et al., 2021
Japan	Kubu river and its tributaries	—	1.3	4.3	4.2	1.5	1.1	3.3	12.0	—	—	—	Nishino et al., 2018
Japan	Tamar river and its tributries	—	<LOQ	4.6	3.1	<LOQ	<LOQ	3.7	7.2	—	—	—	Nishino et al., 2018
Japan	Waste treatment plants	—	3.9	4.8	4.6	<LOQ	0.5	5.5	12.0	—	—	—	Nishino et al., 2018
Canada	Tile drains	50.0	0.0	0.4	0.0	0.0	0.0	0.0	0.0	—	—	—	Schaafsma et al., 2019
Canada	Ditches	119.0	0.0	0.7	0.0	0.0	0.0	0.0	0.0	—	—	—	Schaafsma et al., 2019
China	Yangtze river	120.0	2.5	0.1	4.4	0.3	0.0	1.1	—	—	—	—	Mahai et al., 2019
Vietnam	Lake water from Hannoi	7.0	1.6	<LOQ	2.3	—	—	0.4	—	2.0	1.4	6.4	Wan et al., 2021
USA	River and streams	6.0	1.1	9.1	3.6	—	—	2.4	—	—	—	—	Berens et al., 2021
USA	Lakes	5.0	<LOQ	0.9	1.4	—	—	1.4	—	—	—	—	Berens et al., 2021
USA	Wells	7.0	1.0	1.2	0.3	—	—	0.3	—	—	—	—	Berens et al., 2021
USA	WWTPs	8.0	0.6	5.2	19.0	—	—	0.9	—	—	—	1.4	Berens et al., 2021
Reported maximum neonicotinoid concentrations (ng/L)
Japan	River waters	210.0	23.0	85.0	104.0	—	2.0	202.0	48.0	—	—	—	Sato et al., 2016
Japan	Tap water	21.0	—	7.0	26.0	—	—	—	5.0	—	—	—	Sato et al., 2016
Japan	Rivers in Kubu river and its tributries	—	2.5	6.3	7.0	2.2	1.7	7.9	16.0	—	—	—	Nishino et al., 2018
Japan	Tamar river and its tributaries	—	0.0	47.0	8.4	0.0	0.0	3.7	15.0	—	—	—	Nishino et al., 2018
Japan	Waste treatment plants	—	6.0	6.5	7.3	0.0	0.5	12.0	17.0	—	—	—	Nishino et al., 2018
Japan	River water in Fukui	—	1.2	130.0	55.0	<LOQ	1.2	76.0	270.0	—	—	—	Takeuchi and Nishizawa, 2018
Japan	River systems in Gifu	672.0	2.8	62.2	23.0	0.0	0.0	22.3	239.0	—	—	—	Hayashi et al., 2021
Canada	Tile drains	50.0	1.5	7.0	0.2	0.0	0.0	2.6	0.0	—	—	—	Schaafsma et al., 2019
Canada	Ditches	119.0	0.0	7.2	2.9	0.0	0.1	3.8	0.0	—	—	—	Schaafsma et al., 2019
China**	Yangtze river	120.0	5.2	1.1	16.8	1.4	0.1	11.5	—	—	—	—	Mahai et al., 2019
Vietnam	Lake water from Hannoi	7.0	18.6	<LOQ	5.4	—	—	1.8	—	6.0	6.3	17.3	Wan et al., 2021
Philipines	Surface water	6.0	—	—	3.0	—	—	0.2	—	—	—	—	Bonmatin et al., 2021
Philipines	Surface water	2.0	—	—	5.2	—	—	0.2	—	—	—	—	Bonmatin et al., 2021
USA	WWTPs	8.0	4.7	32.0	48.0	—	—	1.4	—	—	—	17.0	Berens et al., 2021

N; number of subjects, ACE; acetamiprid, CLO; clothianidin, IMI; imidacloprid, NIT; nitenpyram, THI; thiacloprid, THXM; thiamethoxam, DIN; dinotefuran, dn-IMI; desnitro-imidacloprid, IMI-urea; imidacloprid-urea, N-dm-ACE; dm-acetamiprid, <LOQ; below quantification limit.

**  Data represents 95^th percentile value.

Table 2 Regulatory limits for neonicotinoids in Japan

Type of standard	Standard name	Standard description	Imidacloprid	Acetamiprid	Thiacloprid	Thiamethoxam	Nitenpyram	Dinotefuran	Clothianidin
Environmental Standards (ng/L)	Registration withholding**	Aquatic plants and animals	1,900	2,500	3,600	3,500	11,000	12,000	2,800
		Water pollution	150,000	180,000	31,000	47,000	1,400,000	580,000	250,000
	Predicted environmental***	Aquatic plants and animals	4,500	24	450	—	—	7,500	—
		Water pollution	15,000	180	—	14,000	—	27,000	12,000
Food standard (μg/kg)	Acceptable daily intake (ADI)**	—	57	71	12	18	530	220	97
	Acute reference dose (ARfD)**	—	100	100	31	500	600	1,200	600

**  Terayama et al., 2020; 2022, FSC Japan, 2022.

***  Sato et al., 2016.

Table 3 Maximum Residual Limits (MRLs) (mg/kg) of neonicotinoids in tea leaves (adopted from Ikenaka et al., 2018)

	Acetamiprid	Clothianidin	Dinotefuran	Imidacloprid	Nitenpyram	Thiacloprid	Thiamethoxam
Japan	30	50	25	10	10	30	20
US	50**	70****	—	—	—	—	20***
CODEX	—	0.7	—	—	—	—	20
EU	0.1*	0.7	—	0.05*	—	10	20

*  Indicates lower limit of analytical determination.

**  There are no US regulations as of February 10, 2010, for the use of acetamiprid on dried tea leaves.

***  There are no US regulations as of March 27, 2013.

****  No US regulations.

Table 4 Demographic characteristics of the Japanese subjects who participated in neonicotinoid biomonitoring studies from 2014–2023 (N=2,190)

Publication	Subject description	Number of subjects	Male	Female	Age range	Type of sample
Ueyama et al., 2014	Japanese adults	52	41	11	40.9±10.5 years	Spot urine
Harada et al., 2016	Adult, general population	373	45	328	18–87 years	Spot urine
Marfo et al., 2015	Hospital patients,	85	25	60	4–87 years	Spot urine
Osaka et al., 2016	Children	223	108	115	3 years	First morning void
Ikenaka et al., 2019	Children	46	23	23	3–6 years	Spot urine
Ichikawa et al., 2019	PND 1-2	57	36	21	23–34 years	Diaper urine
	PND 14	59	37	22	23–35 years	Diaper urine
Ueyama et al., 2019	Children	50	19	31	3 years	Spot urine
Oya et al., 2021	Infants	1,077	548	529	16–23 months	Diaper urine
Anai et al., 2021	Pregnant women	105	—	105	20–40 years	Spot urine
Nimako et al., 2022	Conventional diet consumers	63	28	35	<1–57 years	Spot urine

* PND means postnatal day; yr means year.

Table 5 Neonicotinoid exposure rates reported in various human populations across the world

Country	Year	Sunject description	Neonicotinoid compounds											Neonicotinoid metabolites										Reference
			N	ACE	IMI	THXM	CLO	THI	DIN	NIT	FLO	IMZ	SUL	UF	DN	N-dm-ACE	5‑OH-IMI	IMI-olefin	6‑CNA	Dn-IMI	Dm-CLO	TA	N‑dm-THXM
Reporter urinary detection frequencies (%)
Japan	2014	Japanese adults	52.00	56.00	96.00	100.00	96.00	67.00	100.00	29.00	—	—	—	—	—	—	—	—	—	—	—	—	—	Ueyama et al., 2014
Japan	2016	Adult, general population	373.00	24.40	76.70	92.00	96.50	7.80	93.30	11.80	—	—	—	—	—	100.00	—	—	—	—	—	—	—	Harada et al., 2016
Japan	2016	Hospital patients	85.00	0.00	0.00	8.20	1.20	1.12	—	4.70	—	—	—	—	—	16.50	—	—	—	—	—	—	—	Marfo et al., 2015
Japan	2016	Children	223.00	12.10	15.20	25.10	8.10	0.00	57.80	20.60	—	—	—	—	—	—	—	—	—	—	—	—	—	Osaka et al., 2016
Japan	2019	Nontoilet trained chidren	50.00	10.00	0.00	24.00	18.00	2.00	84.00	—	—	—	—	—	—	78.00	—	—	—	—	—	—	—	Ueyama et al., 2019
Japan	2019	Infants	65.00	—	—	—	—	—	—	—	—	—	—	—	—	24.60	—	—	—	—	—	—	—	Ichikawa et al., 2019
Japan	2019	Infants	65.00	—	—	—	—	—	—	—	—	—	—	—	—	11.90	—	—	—	—	—	—	—	Ichikawa et al., 2019
Japan	2019	Children (after insecticide spraying)	46.00	11.00	15.00	37.00	52.00	30.00	54.00	30.00	—	—	—	—	—	93.00	—	—	—	—	—	—	—	Ikenaka et al., 2019
Japan	2021	Children	1,036	37.30	40.40	41.10	32.70	21.80	21.80	—	—	—	—	—	—	—	—	—	—	—	—	—	—	Oya et al., 2021
Japan	2021	Pregnant women	109.00	0.60	0.60	83.40	80.90	0.00	10.90	—	—	—	—	—	—	5.80	—	—	—	—	—	—	—	Anai et al., 2021
Japan	2022	Conventional diet consumers	63.00	17.30	47.60	38.10	46.50	3.60	61.90	5.30	—	—	—	—	—	69.90	—	—	—	—	—	—	—	Nimako et al., 2022
Average Japan			197.00	18.70	32.40	49.90	48.00	14.80	60.50	16.90	—	—	—	—	—	50.00	—	—	—	—	—	—	—
China	2019	General population	324.00	96.00	97.00	98.00	99.00	92.00	96.00	—	—	—	—	—	—	—	—	—	—	—	—	—	—	Zhang et al., 2019b
China	2020	General population	129.00	38.00	86.00	90.70	96.97	—	69.00	—	—	—	—	—	—	96.90	—	93.00	—	100.00	—	—	—	Wang et al., 2020a
China	2020	Chinese children	289.00	1.00	3.50	33.20	44.60	0.00	0.70	3.80	—	0.00	—	—	—	62.60	0.00	—	—	—	8.70	—	3.80	Wang et al., 2020b
China	2021	General population	386.00	99.80	100.00	100.00	99.80	99.00	90.50	90.80	88.80	79.30	100.00	—	—	100.00	—	—	—	—	—	—	—	Zhou et al., 2021
China	2021	University students	160.00	93.00	90.00	83.00	92.00	81.00	92.00	—	—	—	—	93.00	—	93.00	98.00	—	—	—	—	—	—	Zhang et al., 2021
China	2021	General population	196.00	67.00	85.00	86.00	75.00	91.00	73.00	—	—	—	—	72.00	—	67.00	—	86.00	—	—	—	—	—	Xu et al., 2021
China	2022	Chinese children	305.00	95.10	95.70	99.70	93.40	94.80	97.70	—	—	—	—	—	—	—	—	—	—	—	—	—	—	Zhao et al., 2022
China	2022	Adult population	114.00	18.00	44.00	80.00	79.00	19.00	87.00	30.00	42.00	—	—	86.00	—	100.00	92.00	—	96.00	—	—	27.00	82.00	Li et al., 2022
China	2022	Pregnant women	408.00	16.90	50.00	57.10	52.00	—	—	—	—	—	—	—	—	99.80	98.50	99.30	—	89.20	79.20	—	—	Mahai et al., 2022
China	2022	Pregnant women	408.00	25.20	78.70	45.30	64.20	—	—	—	—	—	—	—	—	99.30	98.80	99.30	—	88.70	65.40	—		Mahai et al., 2022
China	2022	Pregnant women	408.00	37.00	81.10	70.10	66.90	—	—	—	—	—	—	—	—	100.00	98.30	99.50	—	93.40	82.60	—	—	Mahai et al., 2022
Average China			284.27	53.36	73.73	76.65	78.44	68.11	75.74	41.53	65.40	39.65	100.00	83.67	—	90.96	80.93	95.42	96.00	92.83	58.98	27.00	42.90
Nine countries	2020	General population	566.00	21.20	89.40	72.30	72.10	13.10	66.80	53.70	47.70	69.10	15.20	—	—	97.70	—	—	91.90	—	—	41.90	57.10	Li and Kannan, 2020
Germany	2023	Adult population	39.00	10.30	35.90	2.60	—	—	—	—	—	—	—	—	—	53.80	20.50	17.00	—	—	—	—	—	Wrobel et al., 2023
Brazil	2023	Pregnant women	43.00	0.00	16.30	9.30	—	—	—	—	—	—	—	—	—	60.50	14.00	9.30	—	—	—	—	—	Wrobel et al., 2023
USA	2019	Anonymouse adults	60.00	2.00	30.00	—	37.00	0.00	—	—	—	—	—	—	—	90.00	42.00	—	—	—	—	—	—	Baker et al., 2019
USA	2023	General population	47.00	0.10	0.50	1.00	1.00	0.20	0.50	1.00	0.40	1.00	0.20	—	—	0.90	—	—	1.00	—	—	1.00	0.90	Thompson et al., 2023
USA	2023	General population	44.00	0.10	0.20	1.00	1.00	0.30	0.90	1.00	0.30	1.00	0.10	—	—	0.90	—	—	1.00	—	—	1.00	1.00	Thompson et al., 2023
USA	2023	General population	24.00	0.20	0.10	1.00	1.00	0.10	0.70	1.00	0.30	1.00	0.20	—	—	0.70	—	—	1.00	—	—	1.00	1.00	Thompson et al., 2023
Average USA			43.80	0.60	7.70	1.00	10.00	0.20	0.70	1.00	0.30	1.00	0.20	—	—	23.10	42.00	—	1.00	—	—	1.00	1.00
Reporter blood/serum detection frequencies (%)
Japan		*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*
Saudi Arabia	2020	Patients with osteoarthritis	25.00	8.00	68.00	16.00	12.00	0.00	48.00	0.00	0.00	0.00	—	—	—	84.00	—	—	76.00		0.00	28.00		Li et el., 2020a
China	2021	General population	196.00	67.00	88.00	64.00	91.00	81.00	86.00	—	—	—	—	97.00	—	93.00	—	87.00	—	—		—	—	Xu et al., 2021
China	2022	Patients with osteoporosis	120.00	92.00	97.00	89.00	97.00	91.00	96.00	—	—	—	—	90.00	—	95.00	—	95.00	—	—		—	—	Zhang et al., 2022a
China	2022	Nonosteoporosis population	80.00	93.00	95.00	91.00	98.00	90.00	95.00	—	—	—	—	90.00	—	89.00	—	94.00	—	—		—	—	Zhang et al., 2022a
China	2022	Pregnant women	95.00	0.70	0.70	0.40	0.61	0.70	—	—	—	—	—	—	—	0.70	0.80	—	—	—		—	—	Zhang et al., 2022c
China	2022	Pregnant women	95.00	0.60	0.80	0.50	0.61	0.60	—	—	—	—	—	—	—	0.90	0.80	—	—	—		—	—	Zhang et al., 2022c
China	2022	Healthy population	100.00	61.00	80.00	83.00	94.00	91.00	83.00	—	—	—	—	81.00	86.00	93.00	82.00	85.00	—	—		—	—	Zhang et al., 2022b
China	2022	Patients with liver cancer	27.00	80.00	97.00	61.00	78.00	62.00	80.00	—	—	—	—	86.00	68.00	91.00	75.00	70.00	—	—		—	—	Zhang et al., 2022b
Average China			92.30	50.30	65.80	50.60	58.90	52.00	81.30	0.00	0.00	0.00	—	88.80	77.00	68.30	39.70	86.20	76.00	—	0.00	28.00	—

N; number of subjects, ACE: acetamiprid, IMI: imidacloprid, THXM: thiamethoxam, CLO; clothianidin, THI; thiacloprid, DIN; dinotefuran, NIT; nitenpyram, FLO; flonicamide, IMIZ; imidaclothiz, SUL; sulfoxaflor, UF; 1-methyl-3-(tetrahydro-3-furylmethyl) urea, DN; 1-methyl-3-(tetrahydro-3-furylmethyl) guanidine, N-dm-ACE; N-desmethyl-acetamiorid, 4-OH-IMI; 4-hydroxy-imidacloprid, 5-OH-IMI; 5-hydroxy-imidacloprid, IMI-olefin; Imidacloprid-olefin, 6-CNA; 6-chloronicotinic acid, Dn-IMI; Desnitro-imidacloprid, Dm-CLO; Desmethyl-clothianidin, TA; Thiacloprid-amide, N-dm-CLO; N-desmethyl-clothianidin, N-dm-THXM, N-desmthyl-thiamethoxam.

PESTICIDES: TYPES, USES AND IMPACTS

Pesticides are natural or synthetic chemicals that prevent, destroy, repel, or mitigate pests (Alavanja, 2009; Sharma et al., 2019). Depending on the target pest, pesticides may be classified as insecticides (insect control), rodenticides (rodent control), and herbicides/weedicides (herb/weed control). Other pesticides targeting microorganisms include algicides, fungicides, and bactericides (WHO, 2008; Alavanja, 2009; Sharma et al., 2019). Based on functional class of the active ingredient, insecticides can be further classified as organochlorines, organophosphorus, carbamates, pyrethroids, manganese compounds, and neonicotinoids (WHO, 2008).

The first application of synthetic pesticide was recorded in 1940; since then, pesticide application rates have increased tremendously worldwide (WHO, 2008). The world’s pesticide consumption rate is estimated at 2 million tons per annum, out of which, herbicides, insecticides, fungicides, and other pesticides account for 47.5%, 29.5%, 17.5%, and 5.5%, respectively (Sharma et al., 2019). Financially, the world’s pesticide expenditure totaled approximately ＄56 billion in 2012, with herbicides and insecticides accounting for the largest proportions of the expenditure (44% and 29% (USEPA, 2017)). In terms of geographical distribution, countries such as China, the USA, Argentina, Thailand, Brazil, Italy, France, Canada, Japan, and India are considered the top 10 consumers of pesticides worldwide (Sharma et al., 2019). However, the current pesticide market is growing rapidly in developing countries within Africa, Asia, South and Central America, and the Eastern Mediterranean region (WHO, 2008).

WHAT ARE NEONICOTINOID INSECTICIDES?

Neonicotinoids are nitroguanidine compounds designed with systemic or in-furrow properties. This class of insecticides has several attractive properties, making them appealing for use: for instance, they are (I) predicted to have lower mammalian toxicity due to their relatively higher selectivity for insect nicotinic acetylcholine receptors (nAChRs) than mammalian nAChRs, (II) highly persistent, (III) active against a broad spectrum of crop pests, (IV) versatile in their applications (used for soil and seed treatments, aerial sprays, and foliar treatments), and (V) highly water soluble (Morrissey et al., 2015; Craddock et al., 2019).

Since their introduction, neonicotinoids have become a popular brand of insect control agents, especially in veterinary science and agriculture. They are highly effective against biting, chewing, sucking, and hopping insects. Neonicotinoids are widely used in household, lawn, and garden products to control insects such as ants, cockroaches, and termites. Additionally, they are used in veterinary healthcare as flea and tick preventatives for dogs and cats, in addition to their agricultural uses (Craddock et al., 2019). At present, various neonicotinoid compounds are registered in >120 countries for applications on over 140 types of crops (Chen et al., 2014; Morrissey et al., 2015; Mitchell et al., 2017). Until now, approximately 12 neonicotinoid active compounds have been commercialized for various applications: imidacloprid, acetamiprid, nitenpyram, thiacloprid, paichonding, and cycloxaprid (1^st generation chloropyridyls); thiamethoxam, clothianidin, and imidaclothiz (2^nd generation chlorothiazoles); dinotefuran (3^rd generation furanyls). The newest neonicotinoids to be introduced into the pesticide market include sulfoxaflor (4^th generation sulfoximines) and flupyradifurone (butanolides (Fig. 1)). Among these neonicotinoid compounds, imidacloprid, acetamiprid, clothianidin, dinotefuran, thiamethoxam, nitenpyram, and thiacloprid are most commonly used worldwide (Wang et al., 2017). Over the past 2 decades, neonicotinoid-containing formulations have become important insecticidal brands worldwide (Marfo et al., 2015; Anai et al., 2021). In 2014, the sales of neonicotinoids accounted for over 25% of the global pesticide market, worth more than US ＄3 billion (Craddock et al., 2019).

PLAUSIBLE ENVIRONMENTAL CONTAMINATION PATHWAYS OF NEONICOTINOIDS

Generally, the occurrence of pesticide pollution stern from challenges associated with their bioaccumulation and biomagnification tendencies, persistence, storage, and/or disposal practices (Sharma et al., 2019). Therefore, soil, water, and food may be important pollution media for neonicotinoid pesticides within the ecosystem. Besides, neonicotinoids may contaminate the air due to droplets generated from their aerial applications. Neonicotinoids could be present in unintended food items and linger in the environment (Craddock et al., 2019). Human exposure to neonicotinoid pesticides may occur via spray drifts, house dust, food, and drinking water (Alavanja, 2009; Sharma et al., 2019). Additionally, neonicotinoid residues in agricultural soils may leach into groundwater bodies, posing risks to microbiota and microfauna.

FATE AND KINETICS OF NEONICOTINOIDS IN THE ENVIRONMENT

Once neonicotinoids are applied to the plant seeds, 2%–20% of the active compounds are taken up by the plant; the remaining 80%–98% either seeps into the soil and ground/surface water or is transported through wind drift (Thompson et al., 2020).

In water, neonicotinoid residues may undergo degradation through hydrolysis, biodegradation, and photolysis; however, these processes may be greatly influenced by temperature, salinity, and humic acids (Jia et al., 2023). Under neutral or acidic p^H conditions, neonicotinoid residues resist hydrolysis in water; however, these chemicals degrade rapidly in the water medium in adequate sunlight (Thompson et al., 2020).

Neonicotinoids have low sorption in the soil; they persist in soil with half-life ranging from 1 day to approximately 19 years. Nonetheless, the stability of neonicotinoid compounds in soil can be affected by soil types, moisture, p^H, temperature, and ultraviolet radiation (Thompson et al., 2020). Neonicotinoid residues in soils mainly undergo degradation via microbial processes and photolysis. In the soil degradation studies, biotransformation has been identified as the main pathway for neonicotinoid loss. Most nitro-based neonicotinoids such as thiamethoxam, clothianidin, and imidacloprid, can undergo direct photodegradation while in soil, but the cyano-neonicotinoids, for instance, acetamiprid and thiacloprid, are largely stable in the soil in sunlight (Thompson et al., 2020).

FATE AND KINETICS OF NEONICOTINOIDS IN MAMMALS

Mammalian systems rapidly biodegrade neonicotinoid compounds through metabolic attack at N-heterocyclylmethyl moiety, heterocyclic or acyclic spacers, and N-nitroimine, nitromethylene, or N-cyanoimine tips. Microsomal CYP450 isozymes are essential for phase I metabolism with their selectivity for hydroxylation, desaturation, dealkylation, sulfoxidation, and nitrogen reduction in situ. Some neonicotinoids can be nitro-reduced by cytosolic aldehyde oxidase. Casida (2011) reports that phase II metabolism involves methylation, acetylation, and the formation of glucuronides, glucosides, amino acids, and conjugates derived from sulfate and glutathione. Based on findings from previously published pharmacokinetic studies, the half-lives of neonicotinoids in humans range from 0.23 to 1.45 days (Harada et al., 2016). Meanwhile, due to the low molecular weights and high-water solubility of neonicotinoid compounds, they are primarily excreted from mammalian systems via urine (Thompson et al., 2020).

POTENTIAL TOXICOLOGICAL IMPLICATIONS OF NEONICOTINOIDS INSECTICIDES

Neonicotinoid insecticides mimic the physiological neurotransmitter acetylcholine by agonizing the postsynaptic nicotinic acetylcholine receptors (nAChR), well expressed within the central nervous systems of vertebrates and invertebrates (Morrissey et al., 2015). The agonistic actions of neonicotinoids at nAChR induce continuous excitation of the neuronal membranes, leading to paralysis and lethality in insect species (Simon-Delso et al., 2015; Morrissey et al., 2015). Neonicotinoids are known to show higher selectivity for the nAChR of insects than vertebrates; this has been partly attributed to species-specific differences in the structure and subunits of nAChR (Simon-Delso et al., 2015). As a result of this physiological tendency, neonicotinoids have predicted to elicit minimal toxicological outcomes in mammalian species. In contrast, neonicotinoids in the environment induce intense stress on pollinators and other beneficial insects within the ecosystem, largely due to their high agonistic potencies for insects` nAChR (Siviter and Muth, 2020). Several scientific reports have associated mortalities of beneficial insects with various environmental factors. Among numerous potential environmental stressors, neonicotinoid application has been cited as a major risk factor for the decline of honeybee colonies (Rinkevich et al., 2015). Consequently, the European Union has strongly restricted the application of certain neonicotinoid compounds within the EU region (EFSA, 2013).

Meanwhile, the current dispensation has seen a rise in acute neonicotinoid intoxications in human populations. Several cases of acute neonicotinoid poisoning have been reported worldwide (Phua et al., 2009; Imamura et al., 2010; Lin et al., 2013; Vinod et al., 2015; Casida, 2018); severe human toxicity resulting from deliberate ingestion of neonicotinoid compounds showed neurotoxic manifestations such as status epilepticus, convulsions, and hypotension. Additionally, many mechanistic studies have consistently predicted the potential adverse outcomes of chronic exposures to neonicotinoids in mammalian species. For instance, Li et al. (2011) and Chen et al. (2014) reported that neonicotinoids preferentially distress the α4β2 subtype of the mammalian nAChRs; this process may exacerbate several central nervous system disorders in human systems.

Numerous mechanistic studies have recently suggested that neonicotinoids and their associated metabolites may elicit neurotoxic, hepatotoxic, and reproductive cytotoxic effects in humans through biochemical processes like lipid peroxidation, DNA and protein damage, and oxidative stress-mediated processes (Wang et al., 2017). Moreover, neonicotinoids are known to metabolize into several lipophilic, and hydrophilic compounds within mammalian systems; some of these metabolites are of greater toxicological potencies in mammalian species than the parent neonicotinoid compounds (Wang et al., 2017). For instance, desnitro-imidacloprid (an imidacloprid metabolite) is 300 times more potent to the mammalian nAChRs than the parent compound, imidacloprid (Tomizawa and Casida, 2002; Swenson and Casida, 2013).

CURRENT STATUS OF NEONICOTINOID APPLICATION RATES IN JAPAN

The historical antecedents of pesticide applications in Japan date back to the 1600s, when various mixtures were used predominantly to exterminate pests on agricultural fields (Ota, 2013). Since then, various chemical compounds, including arsenic-based compounds, are used to control pests, and this laid a good foundation for the current robust Japanese pesticide industry. Giant Japanese agrochemical firms such as Mitsui Chemicals, Nippon Soda, and Sumitomo played major roles in neonicotinoid insecticide manufacturing since the early 1990s (Imamura et al., 2010; Craddock et al., 2019). As a result, neonicotinoid-containing formulations are possibly used in Japan since 1992 (Yamamuro et al., 2019); ever since, the domestic application rates of neonicotinoids in Japan have increased exponentially. According to the National Institute for Environmental Studies (NIES 2022), the total shipment amounts of neonicotinoids increased by seven times in 2017 than 1996. In 1992–2001, imidacloprid dominated the number of neonicotinoid active compounds used in domestic pest control systems in Japan; however, from the early 2000s till date, other neonicotinoid compounds such as dinotefuran, clothianidin, and acetamiprid have gained solid ground in the domestic pest control regimen of Japan (Taira, 2014).

According to NIES, an estimated amount of 515.4, 418.9, and 337.9 tons of neonicotinoid compounds were used in Japan in 2016, 2018, and 2020. Within the same period (2016–2020), dinotefuran was Japan’s most locally consumed neonicotinoid compound, followed by clothianidin, imidacloprid, and acetamiprid (Fig. 2A). The average domestic shipment of dinotefuran from 2016 to 2020 exceeded all other neonicotinoids in 83% of 47 Japanese prefectures (Fig. 2B). However, within the Hokkaido and Fukushima prefectures of Japan, the average shipment amount of thiamethoxam from 2016 to 2020 exceeded other neonicotinoids. Similarly, acetamiprid had the highest average shipment value in the Nagano, Yamanashi, and Tokyo, clothianidin had the highest average shipment value in the Ibaraki and Kanagawa, and imidacloprid recorded the highest average shipment value in the Shimane prefectures of Japan, from 2016 to 2020 (Fig. 2B).

STATUS OF NEONICOTINOID CONTAMINATION IN JAPAN ENVIRONMENTS

NEONICOTINOID CONTAMINATION IN JAPANESE AGRICULTURAL SOILS

Soils are the main repository for huge tons of neonicotinoid active compounds used in the Japanese agricultural environments. According to Aseperi et al., (2020), only 5% of the seed-coated neonicotinoids are absorbed by the crop plant, the rest probably remaining in the soil. A previous investigation revealed that soils from agricultural areas might have higher residual concentrations of neonicotinoids than commercial, traffic, residential, industrial, and educational zones and parks (Zhang et al., 2020), suggesting that the soil ecology in agricultural areas could be more susceptible to the exposure and toxicological implications of neonicotinoids. Meanwhile, neonicotinoids degrade quickly in soils under nonsterilized conditions via nitrate reduction, cyano hydrolysis, and chloropyridinyl dichlorination (Zhang et al., 2018). Such soil-neonicotinoid degradation processes may yield chemical products that are much more toxic to soil microorganisms and microbial processes. Depending on the physicochemical conditions of soils, neonicotinoid residues in soils could also leach into groundwater or translocate into soil biological systems. In a previous leaching study, the neonicotinoid thiamethoxam was found to leach out of soils readily (Aseperi et al., 2020). Moreover, clothianidin and thiamethoxam elute readily through sandy soils and pumice (Mörtl et al., 2016). These findings corroborate that the prevalence of neonicotinoid residues in soils could present significant risks of groundwater contamination with possible ecological impacts (Aseperi et al., 2020).

Several reports provided evidence of neonicotinoid infiltrations into soils in countries worldwide. For instance, Wu et al. (2020) detected seven neonicotinoids in soils of tomato greenhouses and six neonicotinoids in soils of cucumber greenhouses in Shouguang, East China (concentrations 0.731–11.383 μg/kg and 0.363–19.224 μg/kg). Yu et al. (2021) determined the residual levels of neonicotinoids in 351 agricultural soil samples from Zengcheng, a typical agricultural zone in South China, and found that at least one neonicotinoid was detected in 95% of the soil samples. In the Philippines, neonicotinoid compounds were detected in 78% of agricultural soil samples (n=67), collected from three provinces, with 0.017–0.89 μg/kg cumulative concentration (Bonmatin et al., 2021). Another study by Bonmatin et al. (2019) found high detection frequencies (Dfs) of neonicotinoids (68%) in soil samples collected from Belize, Central America. Additionally, soils from corn fields in Mongolia were polluted by four neonicotinoids, with thiacloprid being the most predominant in the soil samples (average concentration of thiacloprid of (I) surface soil: 3,901.2±0.04 μg/g and (II) subsurface soil: 3,988.6±0.05 μg/g (Elumalai et al., 2022)).

A critical literature evaluation revealed that neonicotinoid contamination tendencies in agricultural soils might be influenced by factors such as soil type, soil depth, and crop variety. While clayey and loamy soils retained high neonicotinoid concentrations in topsoil, sandy soils eluted neonicotinoids easily into deep soil (Mörtl et al., 2016). Moreover, when neonicotinoid concentrations were measured in greenhouse soils (Wu et al., 2020), the highest concentrations were observed in topsoil samples relative to deep soil samples. Bonmatin et al. (2019; 2021) consistently observed that the distributions of neonicotinoid concentrations in agricultural soils differ among crop types. Soils from vegetable farms contain higher neonicotinoid residues (median; 23.0 ng/g dry weight) than rice paddies (median; 6.1 ng/g dry weight) and fruit farms (median; 5.0 ng/g dry weight) Yu et al., (2021). Such crop-dependent distributions of neonicotinoid residues in agricultural soils have been ascribed to differences in crop planting frequencies, resonating with varying application frequencies of neonicotinoid insecticides.

In Japan, there is no information on the status of neonicotinoid contamination in agricultural soils. Information on the leaching potentials of neonicotinoids in the Japanese soil systems and the status of groundwater contamination by neonicotinoids is blurry. The toxicological implications of the cocktail of neonicotinoid compounds used in Japanese agricultural systems on soil microbiota are not well elucidated.

NEONICOTINOID CONTAMINATION IN JAPANESE WATER SYSTEMS

In Japan, studies have detected neonicotinoid residues in river water systems in various prefectures (Table 1). In the Kanagawa prefecture, Sato et al. (2016) detected imidacloprid, thiamethoxam, clothianidin, dinotefuran, and acetamiprid, with maximum concentrations of 104, 202, 85, 48, and 23 ng/L and Dfs of 67.1%, 22.4%, 53.3%, 30.5%, and 2.4%, respectively, in the river waters. Takeuchi and Nishizawa (2018) also measured residual neonicotinoid concentrations in the Fukui prefecture’s river water bodies and found maximum neonicotinoid concentrations in the following trends: dinotefuran (270 ng/L)>clothianidin (130 ng/L)>thiamethoxam (76 ng/L)>imidacloprid (55 ng/L)>acetamiprid (1.2 ng/L)~thiacloprid (1.2 ng/L)>nitenpyram (<LOQ). In the Tokyo city of Honshu prefecture, the maximum residual concentrations of neonicotinoids reported in the Tama river and its tributaries were as follows: clothianidin (47 ng/L)>dinotefuran (15 ng/L)>imidacloprid (8.4 ng/L)>thiamethoxam (3.70 ng/L)>acetamiprid (<LOQ)>thiacloprid (<LOQ)>nitenpyram (<LOQ) (Nishino et al., 2018). Moreover, some river water bodies within the Gifu prefecture of Japan have been detected with neonicotinoid residues in the following concentration patterns: dinotefuran~median: 11.3 ng/L, max: 239 ng/L; clothianidin~median: 5.6 ng/L, max: 62.3 ng/L; imidacloprid~median: 4.0 ng/L, max: 23 ng/L; thiamethoxam~median: 3.6 ng/L, max: 22.3 ng//L and acetamiprid~median: 2.8 ng/L, max: 2.8 ng/L (Hayashi et al., 2021).

Apart from river water systems, few studies reported neonicotinoid contamination in Japanese tap water and water from waste treatment plants (Table 1). For instance, Sato et al. (2016) detected imidacloprid, clothianidin, and dinotefuran with Dfs of 47.6%, 38.1%, and 9.5%, and maximum concentrations of 26, 7, and 5 ng/L, respectively. Further, water samples from a Japanese waste treatment plant reportedly contain residues of dinotefuran, thiamethoxam, imidacloprid, clothianidin, acetamiprid, and thiacloprid with maximum concentrations of 17.0, 12.0, 7.3, 6.5, 6.0 and 0.52 ng/L, respectively (Nishino et al., 2018).

These reports showed the most predominant neonicotinoids in the Japanese water systems: dinotefuran, imidacloprid, thiamethoxam, and clothianidin. These four neonicotinoids are the major active compounds in most pesticide formulations recommended by many agricultural cooperatives in Japan (Yamamuro et al., 2019; Terayama et al., 2022). Other neonicotinoids such as nitenpyram and thiacloprid were mostly detected at low concentrations in the Japanese rivers because the domestic consumption volume of these compounds is lower than other neonicotinoids (Fig. 2A) (Terayama et al., 2022). Acetamiprid residues were relatively lower in Japanese river waters than other compounds, probably due to its high propensity to metabolize into N-dm-acetamiprid, its main metabolic form.

Until now, residues of various neonicotinoid compounds reported in the water systems of Japan (Table 1) are far below the reference thresholds (registration withholding standards and predicted environmental concentrations (PEC)) set for the protection of aquatic animals and plants and indications of water pollution (Table 2). However, most of the neonicotinoid contamination data reported in Japanese surface waters were limited to the neonicotinoid native compounds, with little or no information on the neonicotinoid metabolites. Given the high propensity of neonicotinoid compounds to undergo degradation under environmental conditions, the scarcity of information on neonicotinoid metabolites in water systems could have attenuated the absolute concentration values of neonicotinoids reported in the Japanese water systems.

By comparing neonicotinoid data across countries (Table 1), the detection rates of dinotefuran and imidacloprid reported in river systems within Kanagawa prefecture of Japan (30.48% and 67%) were seen to be higher the water samples from tile drains (dinotefuran; 2.0% and imidacloprid; 12.0%) and ditches (3.0% and imidacloprid; 13.0%) in Canada. However, the detection rates of clothianidin and thiamethoxam (53.33% and 22.38%) detected in the Japanese river water in Kanagawa were lower than waters from the Canadian tile drains (clothianidin: 88.0%; thiamethoxam: 58.0%) and ditches (clothianidin: 95.0%; thiamethoxam: 50.0%). In addition, Dfs of acetamiprid, imidacloprid, thiamethoxam, thiacloprid, and clothianidin reported in the Yangtze river of China (100%, 100%, 95%, 87%, and 64%) were all higher than the Japanese river systems in Kanagawa prefecture, Japan (Table 1).

Meanwhile, the median concentrations of neonicotinoids, dinotefuran, imidacloprid, and clothianidin, reported in various Japanese rivers (Kubu and Tamar rivers and their tributaries and other river systems in Gifu) were higher than the respective median concentrations of the neonicotinoids reported in rivers, streams, lakes, and wells in the US and Canada (Table 1). Similarly, the median concentrations of neonicotinoids, such as thiamethoxam and clothianidin, in various Japanese rivers and their tributaries were higher than in the Yangtze river of China. Conversely, the median concentrations of imidacloprid reported in water from waste treatment plants in the US (19.00 ng/L) were higher than the Japanese waste treatment plants (4.37 ng/L, Table 1).

NEONICOTINOID CONTAMINATION IN JAPANESE FOOD

In Japan, neonicotinoid insecticides are heavily used in agriculture (Yamamuro et al., 2019). Due to the inherent systemic characteristics of neonicotinoids, low levels of their residues might trickle into locally produced food items. However, few studies attempted to monitor the levels of residual neonicotinoid infiltrations into commercial food products on the Japanese market, mostly focusing on tea. Ikenaka et al. (2018) measured the residual levels of neonicotinoid compounds in green tea leaves and beverages from Japan, and the study detected 7 neonicotinoid parent compounds and 10 neonicotinoid metabolites, with dinotefuran having the highest residual concentration (3,004 ng/g). Two neonicotinoid metabolites, dinotefuran-urea and thiacloprid-amide were most frequently detected in the Japanese tea leaves (92% and 89%). However, the concentrations of all neonicotinoids and their related metabolites detected in tea leaves were below their maximum residue levels (MRLs) in Japan (Table 3). Additionally, Nimako et al. (2022) determined the levels of neonicotinoid remains in organic and conventional Japanese tea leaves purchased from grocery shops in Japan and detected seven neonicotinoid native compounds and one neonicotinoid metabolite in the samples. The detection frequencies of neonicotinoids in the tea samples ranged from 1.94%–84.47%, whereas 94% tea leaves were contaminated with more than one neonicotinoid compound. Dinotefuran, clothianidin, and imidacloprid were detected with the highest concentrations in organic and Japanese green tea samples (8.30, 2.02, and 1.45 ng/g w/w), but the detected concentrations of all compounds were far below their MRLs (Table 3).

These findings are confirmatory that the various neonicotinoid-containing formulations used in the Japanese agricultural fields could trickle into consumable food products for human exposure. However, more information is needed about the residual concentrations of neonicotinoid compounds in commonly consumed food commodities such as fruits, vegetables, rice, etc. This knowledge gap may hamper the implementation of human and ecological risk assessment protocols and the enactment of suitable regulatory frameworks for neonicotinoids.

In contrast, robust data on neonicotinoids has been established for dietary risk assessment systems in other countries. In the US, low levels of neonicotinoids are detected in regularly consumed fruits and vegetables sold on the market. In 1999–2015, residual levels of seven neonicotinoids were consistently measured and reported in various food commodities including fruits, vegetables, meat, dairy, grains, honey, and baby food (Craddock et al., 2019). While the average Df recorded over the entire study period was 4.5%, high Dfs were reported for specific neonicotinoids such as acetamiprid and imidacloprid in various food items such as cherries, apples, pears, and strawberries (45.9%, 29.5%, 24.1%, and 21.3% Df for acetamiprid); cauliflower, cilantro, grapes, kale, lettuce, potatoes, and spinach (57.5%, 30.6%, 28.9%, 31.4%, 45.6%, 31.2%, and 38.7% imidacloprid). Another study by Chen et al. (2014) measured neonicotinoids in fruits and vegetables common to human consumption in Boston, US. Except nectarine and tomato, all fruit and vegetable samples contained at least one neonicotinoid. Moreover, 72% and 45% of fruits and vegetables had two or more neonicotinoid compounds in each sample, and imidacloprid had the highest detection rate. This finding confirmed that agricultural products sold on the open market for human consumption could be contaminated with low-level neonicotinoid residues.

In China, various studies have extensively reported the residual contamination status of neonicotinoids in food commodities. Zhang et al. (2019a) measured and detected at least one neonicotinoid compound in 83 vegetables and 40 fruits purchased from local Chinese markets (supermarkets and conventional markets). The study further observed that green vegetables, baby cabbage, and carrots simultaneously contain six neonicotinoid compounds, and imidacloprid had the highest detection rate in fruits and vegetables (74% and 70%). Another study by Li et al. (2020d) profiled the neonicotinoid residual levels in Chinese tea over 7 years, and out of 726 tea samples tested, 87% were detected with neonicotinoid residues. The imidacloprid concentrations detected in 4.6% of the samples tested were found beyond the Chinese MRLs. Lu et al. (2018) also quantified neonicotinoid residues in 64 fruits and vegetables collected from cafeterias in China and obtained Dfs of 52% and 53% for imidacloprid and thiamethoxam. The observations made from the Chinese studies collectively clarify the ubiquity of neonicotinoid applications in global agriculture and their residual prevalence in dietary staples.

HUMAN EXPOSURES TO NEONICOTINOIDS IN JAPAN

Pesticide exposure in human populations is an important public health concern, mainly due to the propensities of hazardous pesticides to inflict various toxicological deficits in highly susceptible groups within human populations. For instance, children population are highly vulnerable to pesticides exposure, primarily because during the early stages of organ development, pesticide exposures may cause irreversible damage. Children’s vulnerability to pesticides is known to extend from fetal development through infancy, childhood, and adolescence (Tessari et al., 2022). This, in addition to many other factors, warrants the need to mount robust risk assessment protocols for emerging and legacy pesticides in human populations.

The environmental reports, including this one, have triggered comprehensive research into the neonicotinoids exposure dynamics in the Japanese population. Estimatedly, each Japanese adult consumes 1,050, 713, 307, 265, and 206 μg/day of acetamiprid, dinotefuran, imidacloprid, thiamethoxam, and clothianidin, (Harada et al., 2016). Hence, an appraisal of the up-to-date human exposure data of neonicotinoids in the Japanese population could benefit understanding the exposure patterns and potential implications of other emerging pesticides on human populations worldwide.

Based on 10 Japanese biomonitoring publications obtained for this report, a total of approximately 2,190 Japanese subjects participated in neonicotinoid exposure studies from 2014 till date (Ueyama et al., 2014; Harada et al., 2016; Marfo et al., 2015; Osaka et al., 2016; Ichikawa et al., 2019; Ikenaka et al., 2019; Ueyama et al., 2019; Oya et al., 2021; Anai et al., 2021; Nimako et al., 2022 (Table 4)). The studied population comprised 42% males (n=910) and 58% females (n=1,280). These publications also covered a wide range of age groups, including adults (n=530; 24.2% of the total population tested), children aged 3–6 years (n=319; 14.6% of the total population tested), infants below 3 years (n=1,193; 54.5% of the total population tested), and others (n=148; 6.8% of the total population tested). The majority of the Japanese adult population tested for neonicotinoid exposures were females, as the male/female composition ratio of the studied population was 0.19. Children (3–6 years) and infants (<3 years) studied for neonicotinoid exposures in Japan had male/female ratios of 0.89 and 1.09.

Upon intensive data mining and subsequent data evaluations, eight neonicotinoid compounds were frequently detected in the Japanese population: dinotefuran, thiamethoxam, clothianidin, imidacloprid, acetamiprid, N-dm-acetamiprid, nitenpyram, and thiacloprid (Fig. 2). Of these compounds, dinotefuran, N-dm-acetamiprid, thiamethoxam, and clothianidin were the most prevailing in Japanese subjects (average detection frequencies >40% of the total number of subjects who participated in 10 biomonitoring studies published in Japan; Fig. 2). Within the Japanese adult population, the mean detection rates of dinotefuran, N-dm-acetamiprid, thiamethoxam, and clothianidin were consistently >50% (Fig. 3). However, in the Japanese infant population (<3 years), the urinary detection rates of various neonicotinoid compounds were <50% except for dinotefuran, which recorded an average Df of >60% (Fig. 3). In terms of concentration, dinotefuran was consistently detected with the highest urinary concentrations (Fig. 4) in all the human studies on neonicotinoids published in Japan (Ueyama et al., 2014; Harada et al., 2016; Marfo et al., 2015; Osaka et al., 2016; Ikenaka et al., 2019; Ueyama et al., 2019; Oya et al., 2021; Anai et al., 2021; Nimako et al., 2022). In addition, in the most recent biomonitoring reports, the maximum urinary concentrations of dinotefuran, N-dm-acetamiprid, thiamethoxam, and clothianidin were the highest among all neonicotinoid compounds measured (Fig. 4).

Regarding median urinary concentrations, imidacloprid and dinotefuran were the most significantly detected among the neonicotinoids studied in 2014 (Ueyama et al., 2014). In 2016 and 2019, however, Harada et al. (2016) and Ueyama et al. (2019) consistently detected the highest proportion of median urinary dinotefuran concentrations in Japanese subjects (n=373 and n=50, respectively; Table 6). In a recently published paper (Anai et al., 2021; n=105), clothianidin and thiamethoxam were detected with the highest median concentrations among all neonicotinoids measured in a cluster of pregnant women from Japan (Table 6). These findings collectively point out dinotefuran, N-dm-acetamiprid, thiamethoxam, and clothianidin as the main neonicotinoids of most significant human health concerns in the Japanese population. The local consumption tendencies of neonicotinoids within the Japanese population might have instigated these observations. According to NIES, dinotefuran constituted the greatest proportion of the total neonicotinoid formulations shipped to various prefectures in Japan in 2016 2020; this was followed by clothianidin, imidacloprid, and acetamiprid (Fig. 2A).

Table 6 Median serum, blood and urinary concentrations of neonicotinoids reported in various human populations across the world

Country	Year	Sunject description	N	Neonicotinoid compounds										Neonicotinoid metabolites										Reference
				ACE	IMI	THXM	CLO	THI	DIN	NIT	FLO	IMZ	SUL	UF	DN	N‑dm-ACE	5‑OH-IMI	IMI-olefin	6‑CNA	Dn-IMI	Dm-CLO	TA	N‑dm-THXM
Urinary neonicotinoid concentrations reported in μg/g Cre
Japan	1994	Japanese women	20	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	—	—	—	—	—	—	—	—	—	—	—	—	—	Ueyama et al., 2015
Japan	2000	Japanese women	20	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	—	—	—	—	—	—	—	—	—	—	—	—	—	Ueyama et al., 2015
Japan	2003	Japanese women	20	<LOQ	<LOQ	0.23	<LOQ	<LOQ	<LOQ	<LOQ	—	—	—	—	—	—	—	—	—	—	—	—	—	Ueyama et al., 2015
Japan	2009	Japanese women	17	<LOQ	0.39	0.63	<LOQ	<LOQ	0.50	<LOQ	—	—	—	—	—	—	—	—	—	—	—	—	—	Ueyama et al., 2015
Japan	2011	Japanese women	18	<LOQ	<LOQ	0.57	<LOQ	<LOQ	1.80	<LOQ	—	—	—	—	—	—	—	—	—	—	—	—	—	Ueyama et al., 2015
Japan	2016	Adult, general population	373	<LOQ	0.03	0.07	0.27	<LOQ	1.02	<LOQ	—	—	—	—	—	0.4	—	—	—	—	—	—	—	Harada et al., 2016
Japan	2021	Pregnant women	109	<LOQ	<LOQ	7.4	15.3	<LOQ	<LOQ	<LOQ	—	—	—	—	—	<LOQ	—	—	—	—	—	—	—	Anai et al., 2021
Japan	2022	Conventional diet consumers	63	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	0.35	<LOQ	—	—	—	—	—	0.26	—	—	—	—	—	—	—	Nimako et al., 2022
Average Japan			80	<LOQ	0.21	1.78	7.79	<LOQ	0.92	<LOQ	—	—	—	—	—	0.33	—	—	—	—	—	—	—
China	2021	General population	386	2.35	3.42	3.11	5.12	2.24	2.41	3.06	1.81	5.02	1.53	—	—	48.74	—	—	—	—	—	—	—	Zhou et al., 2021
USA	2023	General population	47	0.01	0.10	0.30	0.20	0.01	0.04	0.20	0.01	1.10	0.01	—	—	0.10	—	—	0.90	—	—	0.20	0.10	Thompson et al., 2023
USA	2023	General population	44	0.01	0.01	0.4	0.3	0.01	0.4	0.10	0.01	1.60	0.01	—	—	0.10	—	—	1.00	—	—	0.20	0.20	Thompson et al., 2023
USA	2023	General population	24	0.01	0.01	0.3	0.2	0.01	0.1	0.20	0.01	1.40	0.01	—	—	0.10	—	—	0.90	—	—	0.30	0.20	Thompson et al., 2023
Average USA			38	0.01	0.04	0.33	0.23	0.01	0.18	0.17	0.01	1.37	0.01	—	—	0.10	—	—	0.93	—	—	0.23	0.17
Urinary neonicotinoid concentrations reported in ng/mL
Japan	2014	Japanese adults	52	0.02	1.90	0.50	0.70	0.14	2.30	<LOD	—	—	—	—	—	—	—	—	—	—	—	—	—	Ueyama et al., 2014
Japan	2016	Hospital patients	85	—	—	0.98	1.60	0.14	—	0.52	—	—	—	—	—	1.20	—	—	—	—	—	—	—	Marfo et al., 2015
Japan	2016	Children	223	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	0.44	LOD	—	—	—	—	—	—	—	—	—	—	—	—	—	Osaka et al., 2016
Japan	2019	Children	46	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	—	—	—	—	—	0.39	—	—	—	—	—	—	—	Ikenaka et al., 2019
Japan	2019	Children	46	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	0.14	<LOQ	—	—	—	—	—	0.34	—	—	—	—	—	—	—	Ikenaka et al., 2019
Japan	2019	Infants	65	—	—	—	—	—	—	—	—	—	—	—	—	0.05	—	—	—	—	—	—	—	Ichikawa et al., 2019
Japan	2019	Infants	65	—	—	—	—	—	—	—	—	—	—	—	—	0.09	—	—	—	—	—	—	—	Ichikawa et al., 2019
Japan	2019	Nontoilet trained chidren	50	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	1.99	<LOQ	—	—	—	—	—	0.26	—	—	—	—	—	—	—	Ueyama et al., 2019
Japan	2021	Children	1,036	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	—	—	—	—	—	<LOQ	—	—	—	—	—	—	—	Oya et al., 2021
Average Japan			185	0.02	1.90	0.74	1.15	0.14	1.22	0.52	—	—	—	—	—	0.39	—	—	—	—	—	—	—
China	2019	General population	324	0.01	0.21	0.15	0.24	0.00	0.14	—	—	—	—	—	—	—	—	—	—	—	—	—	—	Zhang et al., 2019b
China	2020	General population	129	<LOQ	0.08	0.21	0.29	—	0.16	—	—	—	—	—	—	0.75	—	0.52	—	0.31	—	—	—	Wang et al., 2020a
China	2020	Chinese children	289	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	<LOQ	—	—	—	1.35	<LOQ	—	—	—	—	—	<LOQ	Wang et al., 2020b
China	2021	General population	196	0.01	0.59	0.07	0.30	0.02	0.26	—	—	—	—	0.78	—	0.35		1.15	—	—	—	—	—	Xu et al., 2021
China	2021	University students	160	0.02	0.09	0.09	0.19	0.02	0.16	—	—	—	—	0.57	—	0.48	2.08	—	—	—	—	—	—	Zhang et al., 2021
China	2022	Chinese children	305	0.01	0.13	0.21	0.19	0.00	1.64	—	—	—	—	—	—	—	—	—	—	—	—	—	—	Zhao et al., 2022
China	2022	Adult population	114	0.05	0.09	0.17	0.26	0.02	0.65	0.10	0.19	—	—	0.96	—	2.26	0.38	—	0.94	—	—	0.02	0.09	Li et al., 2022
China	2022	Pregnant women	408	<LOQ	0.04	0.07	0.04	—	—	—	—	—	—	—	0.20	1.40	0.64	0.43	—	—	0.05	—	—	Mahai et al., 2022
China	2022	Pregnant women	408	<LOQ	0.04	<LOQ	0.05	—	—	—	—	—	—	—	0.09	0.91	0.53	0.43	—	—	0.02	—	—	Mahai et al., 2022
China	2022	Pregnant women	408	<LOQ	0.03	0.05	0.05	—	—	—	—	—	—	—	0.08	0.77	0.4	0.37	—	—	0.06	—	—	Mahai et al., 2022
Average China			274	0.02	0.14	0.13	0.18	0.01	0.50	0.10	0.19	—	—	0.77	0.12	1.03	0.81	0.58	0.94	0.31	0.04	0.02	0.09
USA	2019	Anonymouse adults	60	<LOQ	<LOQ	—	<LOQ	—	—	—	—	—	—	—	—	0.17	<LOD	—	—	—	—	—	—	Baker et al., 2019
USA	2020	General population	19	<LOQ	0.09	0.06	0.19	<LOQ	—	<LOQ	<LOQ	0.18	0.45	—	—	0.27	—	—	0.45	—	—	<LOQ	<LOQ	Li et al., 2020c
Nine countries	2020	General population	566	<LOQ	0.13	0.11	0.11	<LOQ	0.68	0.01	<LOQ	0.17	<LOQ	—	—	0.26	—	—	0.35	—	—	<LOQ	0.04	Li and Kannan, 2020
Germany	2023	Adult population	39	<LOQ	<LOQ	<LOQ	—	—	—	—	—	—	—	—	—	0.38	<LOQ	<LOQ	—	—	—	—	—	Wrobel et al., 2023
Brazil	2023	Pregnant women	43	<LOQ	<LOQ	<LOQ	—	—	—	—	—	—	—	—	—	0.22	<LOQ	<LOQ	—	—	—	—	—	Wrobel et al., 2023
Serum/blood neonicotinoid concentrations reported in ng/mL
Japan	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*	*
Saudi Arabia	2020	Patients with osteoarthritis	25	0.00	0.04	0.06	0.03	—	0.04	—	—	—	0.00	—	—	0.03	—	—	0.02	—	—	0.00	—	Li et el., 2020a
China	2022	Healthy population	100	0.29	0.21	0.19	0.31	0.29	0.62	—	—	—	—	0.73	0.24	0.79	0.36	1.28	—	—	—	—	—	Zhang et al., 2022b
China	2022	Patients with liver cancer	27	0.49	0.56	0.20	0.47	0.21	0.69	—	—	—	—	1.88	0.87	1.61	0.79	2.03	—	—	—	—	—	Zhang et al., 2022b
China	2021	General population	196	0.13	0.29	0.08	0.22	0.08	0.12	—	—	—	—	0.80	—	0.53	—	0.78	—	—	—	—	—	Xu et al., 2021
China	2022	Osteoporosis patients	120	0.44	1.00	0.16	1.40	0.06	0.68	—	—	—	—	1.27	—	5.99		3.08	—	—	—	—	—	Zhang et al., 2022a
China	2022	Nonosteoporosis population	80	0.05	0.76	0.02	0.71	0.01	0.24	—	—	—	—	0.73	—	2.02		1.41	—	—	—	—	—	Zhang et al., 2022a
China	2022	Pregnant women	95	0.00	0.79	<LOQ	0.04	0.12	—	—	—	—	—	—	—	0.08		0.05	—	—	—	—	—	Zhang et al., 2022c
China	2023	Pregnant women	95	0.00	1.84	0.01	0.04	0.11	—	—	—	—	—	—	—	0.13		0.10	—	—	—	—	—	Zhang et al., 2022c

N; number of subjects, ACE: acetamiprid, IMI: imidacloprid, THXM: thiamethoxam, CLO; clothianidin, THI; thiacloprid, DIN; dinotefuran, NIT; nitenpyram, FLO; flonicamide, IMIZ; imidaclothiz, SUL; sulfoxaflor, UF; 1-methyl-3-(tetrahydro-3-furylmethyl) urea, DN; 1-methyl-3-(tetrahydro-3-furylmethyl) guanidine, N-dm-ACE; N-desmethyl-acetamiorid, 4-OH-IMI; 4-hydroxy-imidacloprid, 5-OH-IMI; 5-hydroxy-imidacloprid, IMI-olefin; Imidacloprid-olefin, 6-CNA; 6-chloronicotinic acid, Dn-IMI; Desnitro-imidacloprid, Dm-CLO; Desmethyl-clothianidin, TA; Thiacloprid-amide, N-dm-CLO; N-desmethyl-clothianidin, N-dm-THXM, N-desmthyl-thiamethoxam.

Tables 5 and 6 compare human exposure rates of neonicotinoids across various countries. The average detection rates of neonicotinoids reported in the Japanese population (60.5%, 50.0%, 49.9%, 48.0%, 32.4%, 18.7%, 16.9%, and 14.8% for dinotefuran, N-dm-acetamiprid, thiamethoxam, clothianidin, imidacloprid, acetamiprid, nitenpyram, and thiacloprid, respectively (Table 5)) were observed to be relatively lower than the average neonicotinoid exposure rates reported in the Chinese population (75.74%, 90.96%, 76.65%, 78.44%, 73.73%, 53.36%, 41.53%, and 68.11% for dinotefuran, N-dm-acetamiprid, thiamethoxam, clothianidin, imidacloprid, acetamiprid, nitenpyram, and thiacloprid, respectively (Table 5)). In contrast, the mean of median volumed-based urinary neonicotinoid concentrations reported in the Japanese population (imidacloprid; 1.90 ng/L, thiamethoxam; 0.74 ng/L, clothianidin; 1.15 ng/L, thiacloprid; 0.14 ng/L, dinotefuran; 1.22 ng/L, and nitenpyram; 0.52 ng/L (Table 6)) exceeded those reported in human subjects in China (imidacloprid; 0.14 ng/L, thiamethoxam; 0.13 ng/L, clothianidin; 0.18 ng/L, thiacloprid; 0.01 ng/L, dinotefuran; 0.50 ng/L, and nitenpyram; 0.10 ng/L (Table 6)). Similarly, the mean of median creatinine-adjusted concentrations of key neonicotinoid compounds such as imidacloprid, thiamethoxam, clothianidin, dinotefuran reported in Japan (0.21, 1.78, 7.79, and 0.92 μg/g Cre (Table 6)) were higher than that reported in the US (0.04, 0.33, 0.23 and 0.18 μg/g Cre (Table 6)). Meanwhile, the average neonicotinoid Dfs reported within the Japanese population (60.5%, 50.0%, 49.9%, 48.0%, 32.4%, 18.7%, 16.9%, and 14.8% for dinotefuran, N-dm-acetamiprid, thiamethoxam, clothianidin, imidacloprid, acetamiprid, nitenpyram, and thiacloprid, respectively (Table 5)) were relatively higher than those reported in the US population (0.7%, 23.1%, 1.0%, 10.0%, 7.7%, 0.6%, 1.0%, and 0.2% for dinotefuran, N-dm-acetamiprid, thiamethoxam, clothianidin, imidacloprid, acetamiprid, nitenpyram, and thiacloprid, respectively (Table 5)).

Although the Chinese urinary biomonitoring reports successfully determined neonicotinoid metabolites in the Chinese population, information on human urinary concentrations of neonicotinoid metabolites remains scanty in Japan (Tables 5 and 6). Moreover, various biomonitoring studies have effectively quantified and reported human blood/serum levels of neonicotinoid compounds and neonicotinoid-related metabolites in human populations from China (with average Dfs ranging from 0.00% to 88.8%) and Saudi Arabia (with Dfs ranging from 0.00% to 76.0%), but there is no existing data on human blood levels of neonicotinoids in Japan (Tables 5 and 6).

NEONICOTINOID EXPOSURES IN ECOLOGICAL SPECIES OF JAPAN

Although neonicotinoid insecticides offer numerous benefits to society, uncontrolled applications of these synthetic pesticides may devastate ecosystem diversity. At present, reports of massive consumption rates and high environmental ubiquity of neonicotinoid-containing formulations have sparked several concerns about their potential adverse effects on biodiversity. These concerns are justified as neonicotinoid spillovers or drifts from agricultural areas could contaminate various ecological habitats and pose risks to species’ populations, health, reproduction, and trophic dynamics (Graves et al., 2023).

NEONICOTINOID EXPOSURES IN HONEYBEES

Several bee species encounter interrelating stressors in the environment, including habitat losses, parasite or pathogen infestations (native and introduced), and pesticide poisoning (Goulson et al., 2015). Reportedly, the recent decline in the number of honeybees, Apis mellifera, in colonies threatens crop production and wild plant community biodiversity (Taniguchi et al., 2012). Neonicotinoid contributions to pollinator declines in the US and Europe generated several controversies (Suryanarayanan, 2015; Botías et al., 2015).

In Japan, only a handful of reports on the susceptibility of invertebrate species to neonicotinoids are present. Taniguchi et al.’s (2012) study explored the actual honeybee damage caused by pesticide use by asking Japanese beekeepers to identify bee losses in 2008–2010. The numbers of damaged honeybee hives recorded in the study were 11,659, 11,533, and 8,328 for 2008, 2009, and 2010. These damages corresponded to estimated amounts of ¥201.1, ¥253.8, and ¥178.0 million, respectively. Among the various pesticide classes considered, neonicotinoid formulations were responsible for 91.4% and 81.7% of the damaged hives in 2009 and 2010, corresponding to 93.2% and 92.4% in 2009 and 2010, respectively. These astounding findings established strong evidence, indicating that neonicotinoids could threaten the bee population within Japanese ecosystems. Over the years, most studies concerning bee susceptibility to neonicotinoids have focused primarily on the European species, A. mellifera. However, a study by Yasuda et al., (2017) determined the acute contact toxicity of A. cerana (localized bee species in the Asian region) and A. mellifera to neonicotinoids and other pesticides, including fipronil, organophosphorus, synthetic pyrethroids, etc. First, A. cerana was found to be 8–14 times more sensitive than A. mellifera, connoting that the ecological effects of pesticides could be more severe in A. cerana, compared to other bee species. Meanwhile, A. cerana showed the highest sensitivity to dinotefuran (0.0014 lg/bee), followed by thiamethoxam (0.0024 lg/bee) and fipronil (0.0025 lg/bee). At present, dinotefuran is more extensively used in Japan than other neonicotinoids, potentially creating a substantial hazard to the localized honeybee species within the Japanese ecosystem. Applications of neonicotinoids on crops could contaminate the pollen and nectar of nearby wildflowers, inadvertently serving as the major route of exposure for bees (Botías et al., 2015). An important lesson learned from these findings is that the toxicological implications of neonicotinoids on endemic vertebrate species in Japan could be more severe than what is reported in other species worldwide. Hence, there is an urgent need to establish robust neonicotinoid risk assessment protocols for Japan’s locally endemic invertebrate species.

CURRENT EVIDENCE OF COLONY CONTAMINATION BY NEONICOTINOID INSECTICIDES

Neonicotinoid insecticides are systemic and may translocate through the xylem and phloem of treated plants into pollen and nectar, posing significant risks to honeybees (Van der Sluijs et al., 2013). Over the last decade, reports have consistently confirmed the residual infiltration of neonicotinoids into honeybee colonies worldwide. In 2017, Mitchell et al. (2017) examined 198 honey samples from several nations to evaluate neonicotinoid exposure of pollinators globally. The study found at least one out of five tested compounds (imidacloprid, acetamiprid, thiacloprid, clothianidin, and thiamethoxam) in 75% of all the honey samples; 45% of the samples contained two or more neonicotinoid compounds, and 10% had four or five. Another study by Mulati et al. (2018) detected acetamiprid, thiamethoxam, and imidacloprid residues in pollen and honey, and the detected contamination levels were apiary-dependent. Intriguingly, neonicotinoid thiamethoxam concentrations detected in pollen from apiaries in horticultural farms were more than four times higher than its acceptable MRL in apicultural products. Similarly, Hernando et al. (2018) monitored residual neonicotinoid contamination in honeybee colonies in sunflowers, grown from seeds treated with thiamethoxam or clothianidin precultivation. The study eventually detected clothianidin and thiamethoxam residual concentrations in beebread and adult bees within 0.10–2.89 ng/g and 0.05–0.12 ng/g and 0.10–0.37 ng/g and 0.01–0.05 ng/g, respectively. The findings made by these studies strongly substantiate the global exposure prevalence and vulnerabilities of honeybees to neonicotinoid insecticides.

IMPACTS OF NEONICOTINOID INSECTICIDE CONTAMINATION ON APICULTURE

Since 2007, public attention has been engrossed by reported drastic declines in honeybee colonies, driven by the colony collapse disorder (CCD). In the CCD phenomena, captive honeybees frequently disappeared from their colonies, leaving behind the queen, young bees, honey, and pollen. Ever since this phenomenon was known, the rate of annual hive losses has skyrocketed by approximately 30% (Suryanarayanan, 2013).

The accumulated levels of neonicotinoids in honeybee colonies inadvertently affect the quality and viability of bee colonies, leading to significant losses and bee extermination (Hernando et al., 2018). Most beekeepers blame neonicotinoids for heavy-hive losses in most countries’ ecosystems. After the commercial introduction of neonicotinoids in 1990, the French beekeepers raised serious concerns about the massive colony losses because of imidacloprid exposures (Carreck et al., 2017). Similarly, neonicotinoid exposures accounted for 70% CDD cases reported in Ontario, Canada (Cutler et al., 2014). In Japan, sporadic reports of large numbers of dead honeybees were made across various prefectures in 2005, especially in Iwate, Hokkaido, and Kyushu prefectures, where neonicotinoids, clothianidin, and dinotefuran were used for controlling shield bugs from rice paddies (JEPA, 2023). According to the Ministry of Agriculture, Forestry, and Fisheries’ honeybee study report from April 2010, clothianidin and dinotefuran were found in 92.3% of dead honeybees delivered to beekeepers. Estimatedly, approximately 12 prefectures of Japan are immensely affected by honeybee colony losses as a result of neonicotinoids (JEPA, 2023).

Furthermore, scientific findings suggest that neonicotinoid insecticides have different toxicity levels for honeybees (Iwasa et al., 2004). The scientific community has agreed that a variety of variables, including specific agricultural pesticides, chemicals used by beekeepers, inadequate diet, microbiological infections, and parasites, contribute to fast hive losses (Suryanarayanan, 2013). Hence, further studies are necessary to offer specific information on the involvement of emerging pesticides such as neonicotinoids in CDD.

AVIAN SPECIES: EXPOSURES AND SUSCEPTIBILITY TO NEONICOTINOIDS

Neonicotinoid insecticide applications are strongly associated with recent declines in insectivorous and grassland bird populations, generating serious conservation concerns for many globally endangered species (Graves et al., 2023). Neonicotinoids may elicit adverse effects directly or indirectly. For instance, birds may be indirectly affected by the neonicotinoid-mediated collapse of the invertebrate population, serving as a major food source for the birds (Hallmann et al., 2014; van Lexmond et al., 2015). Conversely, neonicotinoids may directly elicit effects on birds via ingestion of contaminated invertebrates and coated seeds, through water consumption, breathing, or preening after the accumulation of spray droplets on feathers (Humann-Guilleminot et al., 2019; Addy-Orduna et al., 2019).

In Japan, toxicological and/or risk assessment reports of neonicotinoids in avian species are highly scarce. We found only one study, which tested the susceptibility of Japanese insectivorous birds to neonicotinoids by exposing Japanese quail to approximately 3% and 9% of imidacloprid LD₅₀ for Japanese quail, for 1 or 10 days (Bean et al., 2019). Imidacloprid was cleared below detection limits in all tissues of the Japanese quail within 24 h of exposure. The study further observed an extensive imidacloprid metabolism, with 5-OH-imidacloprid and imidacloprid-olefin being the most predominantly detected metabolites in tissues and fecal samples of the quail. In the substantive study, the tissue accumulation potentials of the imidacloprid compounds and their potential toxicological implications were not elucidated. Notably, there is no information on the environmental exposure levels or their associated risks in the diverse avian species within the Japanese ecosystems.

Many studies in other countries have chronicled the field susceptibility of avian species to the deleterious effects of neonicotinoids. Addy-Orduna et al.’s (2019) study investigated the acute toxicities of imidacloprid, clothianidin, and thiamethoxam in the South American eared doves (Zenaida auriculata). Three insecticides induced reductions in food consumption rates, leading to body weight losses. In doves, imidacloprid was 70 times more toxic than clothianidin and thiamethoxam. Furthermore, Humann-Guilleminot et al. (2019) studied the effects of acetamiprid on house sparrows’ sperm quality and oxidative status after administering a low and field-realistic dose of the compound orally. The acetamiprid-treated birds had more significant declines in sperm density than control birds. In the Netherlands, the bird population is reportedly declining by an average of 3.5% due to surface-water contamination by imidacloprid (Hallmann et al., 2014). In the US, Li et al. (2020b) embarked on a large-scale study, into the impacts of neonicotinoid insecticide applications on the biodiversity of avian species. By employing a dataset on breeding birds and pesticide use in the US, the study found a significant decline in bird biodiversity in 2008–2014 in relation to increase in neonicotinoid application rates within the country. The decline was most eminent in grassland and insectivorous birds, which recorded mean annual reduction rates of 4% and 3%, respectively. Graves et al. (2023) detected clothianidin residue in the carcasses of an adult and nestling Tricolored Blackbird (40 and 7 ng/mL) from breeding colonies in Kern County, California, US. These findings collectively suggest that different avian species could be exposed to various neonicotinoids and hence may be susceptible to the adverse outcomes of these compounds.

EXPOSURES AND SUSCEPTIBILITY OF WILD ANIMALS TO NEONICOTINOIDS

Current knowledge on exposure dynamics, susceptibilities, and implications of neonicotinoids on the wildlife population living in agricultural areas is vague. A previous study extensively determined pesticide residues, including neonicotinoids, in meat products from three game species: wild boar (n=42), roe deer (n=79), and deer (n=15) obtained from northeastern Poland (Kaczyński et al., 2021). Approximately 92% of the meat samples from these species contained pesticides, including neonicotinoids, organochlorine compounds, fungicides, and others. Among 28 pesticide compounds detected, 5 were neonicotinoids. Findings from this study shed light on the importance of establishing ecological risk assessment protocols for neonicotinoids in wildlife species.

Information on neonicotinoid exposure in unrestricted wildlife is vital for clarifying their plausible effects on wildlife health; hence, MacDonald et al. (2018) assessed the contamination status of neonicotinoids and other pesticides in liver samples from 40 wild turkeys and compared the detected levels across Southern Ontario, Canada. Approximately 22.5% of the wild turkeys had detectable levels of neonicotinoid residues, mainly clothianidin and thiamethoxam. The maximum concentrations of thiamethoxam and clothianidin detected in the wild turkeys were 0.16 and 0.12 μg/mL, respectively.

A recent study from Japan determined the urinary exposures and cytochrome P450-dependent metabolism of neonicotinoids in wild raccoons captured in Hokkaido, Japan; 90% of the target raccoons (n=59) were detected with neonicotinoids, and the average total urinary concentration of neonicotinoids was 3.1 ng/mL (Shinya et al., 2022). Further, the study proved that wildlife species such as raccoons could be more sensitive to neonicotinoids than rodent species such as rats, primarily due to raccoons’ lower metabolic capacity for neonicotinoids than rats. These findings are provocative for further studies to be launched into the potential adverse effects of neonicotinoids on various wildlife species.

POTENTIAL EFFECTS OF NEONICOTINOID PESTICIDES ON AQUATIC SPECIES

Residual pesticide infiltration into aquatic environments is an important global issue because these chemicals in aquatic environments may present significant adverse outcomes for various species of aquatic biota. By employing zooplankton, water quality, and annual eel and smelled fishery yield, Yamamuro et al. (2019) discovered that the neonicotinoid applications rates since 1993 were associated with an 83% decline in average zooplankton biomass, causing smelled harvest to collapse from 240 to 22 tons in Japan’s Lake Shinji within the Shimane prefecture. This staggering finding suggested that the high toxicity of neonicotinoid insecticides could threaten the Japanese aquatic ecology. For instance, it is reported that the neonicotinoid acetamiprid might have contributed to the dragonfly population decline in the Gifu prefecture of Japan (Karube et al., 2019). Other studies have shown that the presence of neonicotinoid residues within the Japanese aquatic environment might have resulted in the Chironomus Larvae population decline (a food source for dragonflies) and zooplankton count (Hayasaka et al., 2013, Yamamuro et al., 2019). These vulnerabilities of low tropicilevel organisms to neonicotinoids suggest that organisms occupying high trophic levels of the aquatic food chain could be exposed to significant threats from neonicotinoids. However, there is little information about the biomagnification factors of neonicotinoids and various neonicotinoid-related metabolites within aquatic invertebrates, as well as the sensitivity differences of aquatic organisms to neonicotinoid residues in aquatic environments. Moreover, it is unknown whether neonicotinoids have any significant multigenerational effects in aquatic environments or otherwise.

SUMMARY

Pesticides are widely used to safeguard global food security for the increasing human population (Sharma et al., 2019). Typically, pesticides are used in agriculture, veterinary health care, industry/commerce/government institutions, and homes/gardens. In agriculture, consumption rates of herbicides dominate all other usage patterns of pesticides. For instance, in 2012, the agricultural consumption rates of herbicides dominated over all other pesticide application patterns, accounting for approximately 59% of the total consumption volumes (USEPA, 2017). In other sectors, however, the consumption rates of insecticides dominate other pesticides (Alavanja, 2009; USEPA, 2017). For instance, insecticides accounted for approximately 80% and 50% of pesticide usage in the home/garden and industrial/commercial/governmental sectors (USEPA, 2017). This highlights the significance of insecticide-containing formulations in various spheres of human life.

Among various insecticides, neonicotinoids are considered the most famous worldwide, and their demand continues to increase largely due to the constant need for quality agricultural products and high yields. Although using neonicotinoids has numerous benefits, the balance between their significance to society and effect on human and ecological health remains a puzzle. There has been enough evidence that neonicotinoid residues could trickle into human and environmental compartments, including air, water, food, and soil. From these environmental media, neonicotinoid chemicals may mobilize into biological systems or pose significant risks to humans and other nontarget species.

Our literature search suggests that all seven major neonicotinoid formulations, namely, dinotefuran, imidacloprid, acetamiprid, clothianidin, thiamethoxam, thiacloprid, and nitenpyram are currently used in all 47 prefectures of Japan. The total amounts of shipments of these neonicotinoids in Japan were 515.4, 418.9, and 337.9 tons in 2016, 2018, and 2020, respectively. Among these, dinotefuran was Japan’s most locally consumed neonicotinoid compound from 2016 to 2020, followed by clothianidin, imidacloprid, and acetamiprid. As a result of the ubiquitous usage of neonicotinoids in Japan, various studies have reported residual infiltrations of neonicotinoid compounds in Japanese environmental matrices, especially food and water. Two previous studies consistently detected all seven neonicotinoid compounds and some neonicotinoid-associated metabolites in Japanese commercial green tea, but the detected levels were far below neonicotinoid MRLs. Considering the details of the current literature review, three neonicotinoids, dinotefuran, clothianidin, and imidacloprid, were the most commonly detected compounds in Japanese tea leaves. Similarly, these three neonicotinoid compounds and thiamethoxam were the most prevalent neonicotinoid moieties detected in the Japanese surface-water bodies in Tokyo, Kanagawa, and Fukui prefectures.

Besides the reported environmental presence of neonicotinoids in Japan, the literature establishes prima facie evidence, indicating neonicotinoid exposures are eminent in humans, wildlife, and other nontarget species. Specifically, urinary exposure rates of neonicotinoids were extensively reported in various age groups within the Japanese population. Eight neonicotinoid compounds were frequently detected in the Japanese population: dinotefuran, thiamethoxam, clothianidin, imidacloprid, acetamiprid, N-dm-acetamiprid, nitenpyram, and thiacloprid; dinotefuran, N-dm-acetamiprid, thiamethoxam, and clothianidin were the most prevalent compounds in Japanese subjects (average detection frequencies were >40% of the total number of subjects who participated in 10 biomonitoring studies published in Japan). Regarding the ecological risk assessment, the field exposures of invertebrate species (such as bees) and mammals (including avian species and wildlife species) are less reported in Japan. In effect, risk assessment data on neonicotinoids are highly skewed, as most current reports focus on humans with limited information on animal species.

FUTURE PERSPECTIVES

Despite great advancements made in neonicotinoid research in Japan, major gray areas need to be addressed. From an environmental viewpoint, information on the status of neonicotinoid contamination in the agricultural soils of Japan is highly scarce. Essentially, the toxicological implications of the cocktail of neonicotinoid compounds used in Japanese agricultural systems on soil microbiota are poorly elucidated. Future studies need to address these gaps to provide holistic information on the environmental effects of neonicotinoids. Moreover, information on the leaching potentials of neonicotinoids in Japanese soil systems and the status of groundwater contamination by neonicotinoids needs to be clarified.

Neonicotinoid compounds and their associated metabolites may accumulate in various compartments of the aquatic ecology and pose risks to the aquatic biota. There is no literature on the levels of neonicotinoid metabolites in Japanese ground- and surface-water systems. This is potentially an important knowledge gap to address in future studies where neonicotinoid residue in surface waters may easily undergo microbial and/or photo degradation, and their resultant metabolites could affect the health and wellbeing of the aquatic ecology. Besides, the sediment serves as an important sink and sentinel for pollutant accumulation and redissolution and/or resuspension in aquatic ecosystems. As a result, it will be highly interesting to elucidate the status of residual contamination and the partitioning characteristics of neonicotinoids in Japanese aquatic sediments. However, to our knowledge, there are no known reports on neonicotinoids in sediments from Japanese aquatic environments.

Concerning human health risk assessments, much data has been reported on the Japanese population. Upon a comprehensive review of the available data, however, the following knowledge gaps were identified: First, the urinary sampling strategies used for human studies need to be restructured. Almost all the current neonicotinoid exposure studies in Japan employ spot urine sampling strategies in their studies. Given that the majority of the target neonicotinoid compounds have short biological half-lives, it is highly uncertain whether time-specific spot urine sampling strategies will reveal accurate neonicotinoid exposure levels over 24 h or not. Although most of the human studies normalized the measured spot urinary concentrations of neonicotinoids with creatinine, such an approach could be problematic as urinary creatinine excretion rates of an individual could fluctuate depending on demographic factors such as age, sex, height, and muscle mass (Mage et al., 2008). To accurately estimate the daily exposure trends of short-half-lived neonicotinoids, considering 24-h urine sampling strategies or a duplicated sampling approach to reveal the exact relationships of these chemicals with the dietary patterns of the Japanese might be necessary for subsequent studies. Second, most human studies failed to determine the urinary exposure rates of neonicotinoid metabolites in the Japanese population. Third, although various biomonitoring from China and Saudi Arabia effectively reported human blood/serum levels of neonicotinoid compounds in human populations, data on human blood levels of neonicotinoids in Japan is nonexistent. These knowledge gaps may place impediments in accessing the magnitude of risks posed to sensitive groups within the Japanese population and need to be addressed in future human health risk assessment frameworks of neonicotinoids in Japan.

Finally, toxicological and/or risk assessment data on neonicotinoids in avian species, wildlife, and aquatic organisms are highly scarce in Japan. Especially, there is scanty information on the environmental exposure levels and their associated risks for the diverse avian species within the Japanese ecosystems. Moreover, the sensitivities of aquatic organisms to neonicotinoids are less studied in Japan. More worryingly, it is currently unknown if neonicotinoids could represent any significant levels of multigenerational effects in aquatic environments or otherwise. Perhaps, addressing these knowledge gaps may provide information on the exact impacts of neonicotinoid insecticides on human and animal health and environmental quality.

ACKNOWLEDGMENTS

This research was supported by JST/JICA SATREPS (Science and Technology Research Partnership for Sustainable Development; grant number JPMJSA1501), JST aXis (Accelerating Social Implementation for SDGs Achievement; grant number JPMJAS2001) and AJ-CORE, JSPS Core-to-Core and Hokkaido University (Sosei Tokutei Research).

CONFLICTS OF INTEREST

The authors declare no conflict of interest in the present study.

AVAILABILITY OF DATA AND MATERIALS

Data, associated metadata, and calculation tools are available by contacting the corresponding author ([e-mail]).

HUMAN AND ANIMAL EXPERIMENTATION

Not applicable.

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