Heavy Metal and Antimicrobial Residue Levels in Various Types of Digestate from Biogas Plants—A Review

Abstract

Global population growth generates problems relating to increasing demand for sustainable energy and waste treatment. Proper solid waste management promotes material reuse, maximizes recovery and reduces anthropological pressure on natural resources. Anaerobic digestion (AD) is an alternative method of stabilizing organic substrates and generating biogas as a source of environmentally friendly energy. In addition, digestate is not only a waste product of that process but also a renewable resource with many potential applications. The circular economy concept encourages the use of digestate as a source of nutrients that promotes plant growth and improves soil properties. However, the stabilized substrates often contain various contaminants, including heavy metals (HMs) and antibiotics that are also detected in digestate. Therefore, the agricultural use of digestate obtained by AD could increase the pool of these pollutants in soil and water environments and contribute to their circulation in these ecosystems. Moreover, digestate may also increase the co-selection of genes determining resistance to HMs and antibiotics in environmental microorganisms. This article comprehensively reviews published data on the residues of various HMs and antimicrobial substances in different digestates around the world and maps the scope of the problem. Moreover, the potential risk of residual levels of these contaminants in digestate has also been evaluated. The review highlights the lack of legal standards regulating the concentrations of drugs introduced into the soil with digestate. The results of the ecological risk assessment indicate that the presence of medically important antimicrobials in digestate products, especially those used in agriculture, should be limited.

1. Introduction

Due to the intensification of agriculture, progressing urbanization, and human population growth, the volumes of waste generated worldwide have become a significant burden on the natural environment [1]. The global production of various types of organic waste, including agricultural residues, animal manure, food waste, and sewage sludge, is estimated at 2, 120, 1.3, and 16.4 billion tons per year, respectively [2]. The global volume of organic waste, as well as the consumption and demand for energy, continues to increase. Organic biomass is biodegradable and can be used for energy generation under anaerobic conditions. During anaerobic digestion (AD), various organic wastes are converted into biogas—an environmentally friendly and renewable energy source that can be used to produce vehicle fuel, electricity, or heat. In recent years, AD of waste from agriculture, industry, and municipal facilities has become one of the most promising pathways for renewable energy generation [3]. Sustainable management of organic waste plays an important role in the circular economy by promoting material reuse and maximizing recovery. Energy recovery from organic waste provides environmental benefits by decreasing the following: (1) greenhouse gas emissions; (2) fossil fuel use; (3) water pollution; and (4) the volume of landfilled waste. Moreover, the AD process offers an additional advantage by promoting the simultaneous recovery of material in the form of digestate [4,5]. While biogas as a renewable energy source is discussed intensively, digestate still plays a minor role in political and scientific debates.

Digestate is a stable and valuable by-product of the AD process that is rich in nutrients. It includes a balanced mixture of micro- and macronutrients which are essential for plant growth [6]. Digestate generated by AD can be returned to farms that supply substrates to biogas plants, sold on the market or recirculated [7]. The use of digestate as fertilizer or soil improver offers one of the simplest solutions for managing this by-product, improving the economic sustainability of biogas production and minimizing its negative environmental impacts. Digestate obtained by AD is more stable, hygienic, and abundant in bioavailable nitrogen (in the form of N-NH₄) than undigested organic biomass. It is suitable for agricultural use due to specific properties that can improve the chemical and physical properties of soil, including soil structure, moisture retention, microbial activity, and organic carbon content. Moreover, the recovery of nutrients (such as nitrogen and phosphorus) limits soil erosion. Ultimately, digestate enhances soil quality and fertility. The agricultural use of digestate offers an excellent alternative to reducing the use of chemical fertilizers [5,8].

However, despite these advantages, the introduction of digestate into the soil environment entails some risks. In Europe, agricultural residues and animal manure are the dominant feedstocks for AD. Due to various waste management options and the legal restrictions related to the climate and energy policy, all types of organic waste and by-products are being increasingly used as substrates in biogas production. For this reason, sewage sludge, biowaste, and industrial waste are also commonly processed by AD [6]. It should be emphasized that the specific properties of the obtained digestate are influenced mainly by the type of substrates used in the AD process. Digestate composition and quality are determined by the composition and quality of the feedstock, as well as the efficiency of the AD process [9]. The substrate should be free of contaminants to prevent undesirable substances from reaching the digestate. There are concerns regarding the presence of hazardous compounds, especially heavy metals (HMs) and antibiotics, in digestates obtained from animal, household, or industrial wastes [10,11,12]. These types of pollutants are most widely encountered in animal manure and sewage sludge [13].

Heavy metals are widely distributed in wastewater and animal feed, and they are present in sewage sludge and manure, respectively [14,15,16,17]. In addition, these micropollutants are also detected in the waste stream [18]. Heavy metals are toxic and non-biodegradable, and they pose a serious environmental risk. Antibiotics also undermine the quality of digestate. According to estimates, up to 80% of ingested drugs can be excreted with feces and urine in unchanged form. The antibiotics used in human medicine reach wastewater treatment plants and accumulate in sewage sludge, while veterinary drugs have been detected in manure and slurry [1]. Antimicrobial substances reach anaerobic bioreactors and are released into the environment, posing a threat to the soil, water ecosystems, and plants. Even low concentrations of HMs and antibiotics in various ecosystems can have adverse consequences for the environment [12,19]. The agricultural use of digestate containing HMs and antibiotics leads to the accumulation of these contaminants in the soil, posing a risk for both the environment and public health [1,19,20]. Moreover, the use of contaminated digestate in agricultural land and crop production can exert selective pressure on bacteria, lead to the spread of HMs and promote antibiotic resistance in microorganisms [1,21]. It should be emphasized that, in Europe, the quality and potential applications of digestate are regulated by legal acts pertaining to fertilizers, waste, and soil protection, or a combination of these laws. Although HM concentrations are strictly regulated in digestate that is intended for use as fertilizer or soil improver, statutory limits on drug concentrations have not been introduced to date [13]. The lack of appropriate risk assessment for the safe management of digestate has also been highlighted by other authors [4].

The main objective of this study was to review the current knowledge on the presence of HMs and antibiotics in digestate derived from various feedstocks. The article describes the key reservoirs of these micropollutants and the role of the AD process in their release to the environment. The growing levels of microbial resistance to HMs and antibiotics pose a serious public health concern and an environmental issue worldwide, which is why the potential effects of micropollutant residues in digestate have also been discussed. An ecological risk assessment of HMs and drugs has been performed to determine the severity of the risk resulting from the introduction of digestate into the soil environment. However, the current study has several limitations. Firstly, despite a large number of scientific papers on the presence of HMs and antibiotics in potential AD substrates, including raw animal manure, agricultural residues, sewage sludge, and the organic fraction of municipal solid waste or mixed wastes [22,23,24,25,26,27,28,29,30,31,32], the concentrations of these micropollutants in the resulting digestates have been analyzed by very few authors. Secondly, most studies investigating the fate of HMs and antibiotics during AD included feedstock supplementation before the process, making it impossible to assess the real, environmental concentrations of these pollutants in digestates [33,34,35,36,37]. Moreover, many scientific papers examining the impact of soil fertilization with digestate focused only on the concentrations of HMs and antibiotics in soil samples, while disregarding AD by-products [38,39,40]. In addition, the solid−liquid separation of AD products has also received little attention in the existing literature. To the best of our knowledge, there is a general scarcity of comprehensive review studies addressing the levels of HM and antibiotic residues in various anaerobically stabilized organic wastes. The number of reviewed papers from various regions of the world is presented in Figure 1. The authors were unable to find any studies presenting data on the content of HMs and/or antibiotics in digestates from Africa and South America. The only published studies addressing digestate contamination on the Asian continent came from China (11 articles). This issue was most often studied by research teams from European countries (22 articles), but many of these articles present selective data. Residual contamination is a broad and complex area of research involving different types of feedstocks, AD conditions, digestate fractions, and their environmental impacts, which is why further research is needed to fill these knowledge gaps.

2. Materials and Methods

2.1. Data Sources

In compliance with the Preferred Reporting Items for Systematic reviews and Meta-Analyses (PRISMA) guidelines [41], the articles were selected according to the following four criteria: (i) identification; (ii) screening studies; (iii) eligibility; and (iv) inclusion. PubMed and Google Scholar scientific literature databases were surveyed to find papers published between 1 January 2004 and 30 November 2024.

2.2. Search Strategy

The search strategy is presented in Figure S1 in Supplementary Materials. The following keywords were used in the search strategy: ((anaerobic digestion) AND (digestate) AND ((antibiotic) OR (antimicrobial) OR (drug) OR (heavy metal)) AND ((residue) OR (degradation) OR (concentration) OR (content))) AND ((anaerobic digestion) AND (digestate) AND ((antibiotic) OR (antimicrobial) OR (drug) OR (heavy metal)) AND ((residue) OR (degradation) OR (concentration) OR (content))). These keywords were tailored to each database.

A preliminary search of published scientific articles relating to the subject of this review was conducted to identify the keywords for the advanced search process. The selected keywords are presented in Figure 2. In addition, a reference list of articles was checked manually to find adequate scientific publications for the review. After filtering the literature, 131 scientific publications were selected for this review article.

2.3. Ecological Risk Assessment

The indicators related to the levels of pollution with HMs and pharmaceuticals were computed to evaluate the influence of digestates on the soil ecosystem and the associated hazards. The geo-accumulation index and ecological risk factors were calculated for HM contamination of digestates. The risk quotient (RQ) method was used to assess the effect of pharmaceutical contaminants in digestates on the soil ecosystem, including co-selection for antimicrobial resistance.

2.3.1. Geo-Accumulation Index

The geo-accumulation index (I_geo) measures the level of pollution with inorganic or organic trace substances and the presence of bioelements in sediment or soil. This index is commonly used to assess HM pollution, and it was calculated by comparing the concentrations of HMs in samples with the natural background levels of metals in soils with the use of the following equation [42]:

$$I_{geo}={\log }_2\left({\displaystyle \frac{C_n}{1.5B_n}}\right),$$

(1)

where

C_n is the concentration of the analyzed heavy metal in digestate (mg∙kg⁻¹);

B_n is the geochemical background value of the analyzed heavy metal in soil (mg∙kg⁻¹). Background values were derived from the elemental concentrations in the upper continental crust [43].

The values of the I_geo index were classified on a 5-point scale: <0, practically unpolluted; 0–1, unpolluted to moderately polluted; 1–2, moderately polluted; 2–3, moderately to strongly polluted; 3–4, strongly polluted; 4–5, strongly to extremely polluted; and >5, extremely polluted.

2.3.2. Potential Ecological Risk Index

The ecological risk index (E_r) was applied to evaluate the ecological risk posed by each HM individually. This index was calculated using the following formula [44]:

$$E_r=T_r·C_f,$$

(2)

where

T_r is the toxicity factor of the analyzed HM; the following values were adopted: cadmium = 30, chromium = 2, mercury = 40, nickel = 5, lead = 5, zinc = 1, copper = 5, arsenic = 10, iron = 1, manganese = 1, cobalt = 5, molybdenum = 15;

C_f is the pollution factor.

The toxic response factors for manganese (Mn) and cobalt (Co) were not provided by [44]; therefore, the present study relied on the relevant values computed by other researchers [45,46,47]. The T_r values for silver (Ag) and aluminum (Al) were not available; therefore, the E_r values for these elements were not calculated.

The contamination factor (C_f) is the ratio of metal contamination levels to pre-industrial levels that are commonly found in the Earth’s crust. This factor was calculated using the following formula:

$$C_f={\displaystyle \frac{C_n}{C_b}},$$

(3)

where

C_n is the concentration of the analyzed HM in digestate (mg∙kg⁻¹);

C_ob is the concentration of the analyzed HM in the soil (mg∙kg⁻¹). Background values were derived from the elemental concentrations found in the upper continental crust [43].

The E_r values were classified on the following scale: very high risk (E_r > 320), high risk (160 < E_r < 320), considerable risk (80 < E_r < 160), moderate risk (40 < E_r < 80), and low risk (E_r < 40).

2.3.3. Predicted Environmental Concentrations (PECs)

To assess the influence of antibiotic residues in digestates on the soil ecosystem, the predicted environmental concentrations (PECs) of antibiotics in fertilizer-amended soil (PEC_(soil)) were calculated using the prescribed methodology [48] and the following equation:

$${PEC}_{\left(soil\right)}={Co}_{\left(soil\right)}+{\displaystyle \frac{{MEC}_{\left(digestate\right)}·{APP}_{\left(digestate\right)}}{{DEPTH}_{\left(soil\right)}·{RHO}_{\left(soil\right)}}}$$

(4)

where

Co_(soil) is the background concentration of the analyzed antibiotic in soil before the application of digestate (assumed to be zero in this study);

MEC_(digestate) is the concentration of the analyzed antibiotic in digestate (μg kg⁻¹);

APP _(digestate) is the typical rate of sludge application in soil, usually 0.5 kg m⁻² for agricultural applications [49];

DEPTH_(soil) is the mixing depth, usually set at 0.2 m for agricultural soil;

RHO_(soil) is the bulk density of wet soil (1700 kg m⁻³) [49].

The calculated PEC_(soil) values were compared with the known predicted no-effect concentrations (PNECs) for resistance selection in the environment (PNEC_MIC) and ecotoxicity (PNEC_ENV) used in antibiotic environmental risk assessments for individual antibiotics detected in digestates. This analysis involved only antimicrobial substances with known values of PNEC_MIC/PNEC_ENV [50,51].

The risk assessment based on PNEC_MIC/PNEC_ENV values was conducted with the use of the RQ method according to the following formula:

$${RQ}_{mic/env}={\displaystyle \frac{{PEC}_{\left(soil\right)}}{{PNEC}_{MIC/ENV}}},$$

(5)

RQ_env values were classified on the following scale: RQ > 1—high potential risk for soil-living organisms; 0.1 < RQ < 1—moderate potential risk; and RQ < 0.1—low potential risk. PEC_(soil) values exceeding the PNEC_MIC for specific antibiotics (RQ_mic > 1) were regarded as contributors to selection for antibiotic resistance in soil.

3. Results

3.1. Heavy Metal Content of Various Digestates

The use of digestate as soil fertilizer or improver is an environmentally friendly and cost-effective method of managing AD by-products that enhances soil capability for agriculture, supplies plants with nutrients and promotes waste recycling. However, HM concentrations in various digestates raise serious concerns. Although AD has the potential to reduce microbial loads and antibiotics in the processed substrate, it is not effective in eliminating HMs that are highly stable [52]. Heavy metals can be introduced into the soil when agricultural land is amended with contaminated fertilizers. Although some HMs, including copper (Cu) and zinc (Zn), are essential for the growth of various organisms, high concentrations of these elements can be toxic and can cause damage to nucleic acids and the cell membrane [53,54].

Heavy metal concentrations can differ considerably in various digestates. The origin of green waste or biowaste feedstocks is a significant consideration because aerial deposition of HMs is higher in urban than in rural environments. Moreover, the season of substrate collection also plays an important role due to the peak deposition of HMs in winter [55,56]. However, these variations are related mainly to the type of feedstock processed by AD. In intensive livestock production systems, HMs (mainly Cu, Zn, and arsenic (As)) are used as feed additives to prevent disease, increase weight gains or boost egg production in poultry farming [57]. Even more than 80% of dietary Zn and Cu is excreted in active form by humans and animals, which leads to the accumulation of these HMs in various feedstocks that are processed by AD [24]. The above leads to an increase in the HM content of animal by-products, which is stabilized during the AD process. Animal manure is recognized as a valuable fertilizer, but it is also the main source of HMs [57,58]. Sewage sludge, yet another popular substrate for AD, is an equally important source of HMs [52]. The HMs that are most frequently identified in sewage sludge include Cu, Zn, Co, iron (Fe), chromium (Cr), lead (Pb), mercury (Hg), nickel (Ni), cadmium (Cd), and the highly toxic As and selenium (Se). Heavy metals can occur in sewage sludge in dissolved or precipitated form, and they can also be associated with solid particles [52]. Industrial development increases the concentrations of HMs in wastewater, directly leading to sewage sludge contamination [59]. Surface runoffs and industrial wastewater are the main sources of these micropollutants in sewage sludge [60]. It should be noted that HM concentrations in sewage sludge rarely exceed the regulatory limits, but prolonged application of sewage sludge in agriculture contributes to the accumulation of HMs in soil [61,62,63]. Soil fertilization with both digested manure and sewage sludge increases HM levels in soil, which poses serious environmental risks. As previously mentioned, HM concentrations in raw substrates for AD and in fertilized soil have been extensively researched. However, very few studies have examined the HM content of digestates. The concentrations of HMs in the by-products of AD of various feedstocks are presented in

Table 1. Heavy metal concentrations in digestates derived from various feedstock sources.

Substrate for AD	HM	Concentration	Unit of Measure	Country	References
Mixed cattle manure and maize silage	Cd	0.9–1.6	mg∙kg−1 DM	Poland	[11]
	Cr	26–40
	Hg	0.31–0.49
	Ni	12–19
	Pb	7.9–20
	Zn	595–790
Mixed cattle manure, bovine slurry, and maize silage	Cd	0.2–0.8	mg∙kg−1 DM	Poland	[64]
	Cr	1.5–6.5
	Hg	0.01–0.023
	Ni	5–15
	Pb	2–11.5
Mixed food waste, industrial waste, animal by-products/ manure, and slaughterhouse waste	Cd	0.34–0.37	mg∙kg−1 DM	Sweden	[65]
	Cr	8.2–20
	Cu	41–100
	Hg	<0.1
	Ni	7.2–15
	Pb	2–7.4
	Zn	180–540
Mixed maize silage, slaughterhouse waste, potato pulp, and confectionery press cake	Cd	0.02–0.03	mg∙kg−1 DM	Poland	[11]
	Cr	0.2–0.25
	Hg	0.0037–0.0097
	Ni	0.32–0.37
	Zn	10.4–11.9
Mixed plant residues and cattle manure	Cd	135.64	mg∙kg−1	China	[66]
Mixed sewage sludge and food by-products	As	12.15–18.83	mg∙kg−1	Poland	[67]
	Cd	7.57–10.79
	Cr	143.65–217.47
	Cu	1387–2059
	Fe	4629–6685
	Hg	19.62–37.16
	Mn	124.78–180.53
	Ni	56.9–81.38
	Pb	18.6–63.83
	Zn	3689–5530
Mixed sewage sludge and green waste	Cu	501	mg∙kg−1 DM	France	[68]
	Fe	10,400
	Zn	842
Mixed textile dyeing sludge and soybean okara by-products	Ni	11.23	mg∙kg−1 VS	China	[69]
	Cr	316.97
	Cu	37.47
	Zn	355.7
Municipal organic waste	Cd	1.4	mg∙kg−1 DM	Germany	[70]
	Cr	26.6
	Cu	49.1
	Ni	12.5
	Pb	123.6
	Zn	327.3
Fine fraction of municipal waste	Cu	136	mg∙kg−1 DM	France	[68]
	Fe	12,100
	Zn	634
Organic output of mechanical-biological waste treatment facilities	Cd	0.77–3.5 (SF) <0.58 (LF)	mg∙kg−1 DM	Italy	[71]
	Cr	41.17–77.54 (SF) 2.31–21.78 (LF)
	Cu	255.35–315.77 (SF) 1.91–57.58 (LF)
	Ni	22.61–35.40 (SF) 7.81–32.49 (LF)
	Pb	34.45–344.13 (SF) 0.26–45.10 (LF)
	Zn	811.5–920.28 (SF) 14.28–224.63 (LF)
Pharmaceutical-derived organic wastes	As	0.19	mg∙kg−1 DM	Italy	[72]
	Cd	<0.2
	Cr	<0.5
	Cu	23.4
	Hg	0.41
	Ni	<0.5
	Pb	<1
	Zn	117.2
Poultry manure	Ag	0.15	mg∙kg−1	China	[73]
	As	106.65
	Cd	0.39
	Cr	61.2
	Cu	71.72
	Pb	22.97
	Zn	370.24
Rice straw	Cd	0.4–0.7	mg L−1	China	[74]
	Cu	0.05–0.15
	Pb	0.03–0.1
	Zn	0.15–0.37
Sewage sludge	Al	3262.3	mg∙kg−1	China	[75]
	Fe	2017.8
	Hg	6.37			[76]
	Cu	487	mg∙kg−1 DM	France	[68]
	Fe	9840
	Zn	988
	Al	6270–8025		Poland	[77]
	Cd	2.3–5.4
	Co	3.8–5.9
	Cr	14–17.9
	Cu	87.8–91.3
	Fe	4001–4548
	Mn	241–267
	Mo	3–3.9
	Ni	6.6–8
	Pb	18.8–25.5
	Zn	732–789
	As	14.11	mg∙kg−1		[67]
	Cd	8.09
	Cr	156.58
	Cu	1,496.7
	Fe	4836
	Hg	23.07
	Mn	132.51
	Ni	61.67
	Pb	42.38
	Zn	4303
	Cd	0.9–1.6	mg∙kg−1 DM		[11]
	Cr	28–44
	Hg	0.41–0.51
	Ni	11–17
	Pb	7.13–18
	Zn	645–830
Pig manure	Cu	788	mg∙kg−1 DM	France	[68]
	Fe	3,180
	Zn	842
	As	7.49	mg∙kg−1 VS	China	[78]
	Cd	0.31
	Cu	781.33
	Ni	23.13
	Pb	7.02
	Zn	2146.32
Textile dyeing sludge	Ni	21.74	mg∙kg−1 VS	China	[69]
	Cr	525.92
	Cu	57.46
	Zn	473.08

DM—dry matter; FW—fresh weight; VSs—volatile solids; LF—liquid fraction; SF—solid fraction; Al—aluminum; As—arsenic; Cd—cadmium; Co—cobalt; Cr—chromium; Cu—copper; Fe—iron; Hg—mercury; Ni—nickel; Pb—lead; Se—selenium; Zn—zinc.

.

According to the literature, digestates are most abundant in Zn, followed by Cu, Mn, and As [79]. Although the content of As, Cd, Cr, and Pb in various digestates is much lower relative to the commonly present Cu and Zn, these HMs are regarded as hazardous and highly persistent microelements [80]. The concentrations of HMs in digestates often exceed the values noted in the feedstock. Several studies reported on increasing concentrations of HMs after anaerobic treatment [64,81,82]. The above can be attributed to the production of methane during the AD of organic matter. This process leads to a decrease in the weight and volume of the substrate [83,84], which ultimately increases HM concentrations. Heavy metals are often closely bound to insoluble solids, which is why their concentrations are considerably higher in solid fractions after anaerobic treatment [82]. These micropollutants are present in the liquid fraction, and they are deposited and adsorbed by solid particles or colloids to form precipitates at the end of the treatment [78]. In addition, HM solubilization may be limited under alkaline conditions [71]. Due to the complexity of AD, HMs can be involved in many physical and/or chemical processes, including adsorption on the solid fraction and precipitation as carbonates, sulfides, and hydroxides [85]. The observed increase in HM concentrations in digestate indicates that the impact of this by-product on the soil environment should be examined, especially if digestate is used as a crop fertilizer. It should be emphasized that HM residues in soil, even at low concentrations, can be absorbed by roots and can accumulate in edible plant parts [52,86].

Digestate is a significant source of potential environmental contamination, which is why permissible HM levels for biomass incorporated into the soil are set by national and international laws. In Europe, the relevant limits have been prescribed by the EU Fertilizing Products Regulation (2019/1009). This regulation classifies the materials introduced into the soil into various categories. Digestate has been classified in four categories as organic fertilizer, soil improver, growing medium, and plant biostimulant. The permissible limits of HMs in each of these categories are presented in

Table 2. Permissible levels of HMs in digestates classified as fertilizers (EU 2019/1009).

HM	Unit of Measure	Types of Digestate Used as Fertilizers
		Organic Fertilizer	Organic Soil Improver	Growing Medium	Plant Biostimulant
Cd	mg∙kg−1 DM	1.5	2	1.5	1.5
Cr		2	2	2	2
Hg		1	1	1	1
Ni		50	50	50	50
Pb		120	120	120	120
As		40	40	40	40
Cu		300	300	200	600
Zn		800	800	500	1500

DM—dry matter; As—arsenic; Cd—cadmium; Cr—chromium; Cu—copper; Hg—mercury; Ni—nickel; Pb—lead; Zn—zinc.

. The EU Fertilizing Products Regulation (EU 2019/1009, Table 2) lists eight HMs, and some of the AD by-products listed in Table 1 do not meet the European agricultural standards for fertilizers. In some studies, Ni, Zn, and Cu concentrations in the examined digestates, in particular digested animal manure, significantly exceeded the regulatory limits. The reported HM content of various digestates further highlights the environmental risks associated with the accumulation of these micropollutants in soils. However, some authors found that the application of digestate derived from animal manure and/or plant material can reduce the mobility and bioavailability of HM in soil through complexation, adsorption, and precipitation [28].

3.2. Antimicrobial Residue Levels in Digestates

Intensive use of antimicrobial substances in agriculture, industry, human, and veterinary medicine has led to an increase in the concentrations of antibiotic residues in various types of organic wastes that may be applied as feedstocks for AD [24,87]. The continuous release of antibiotics into the environment and their accumulation in organic biomass pose a great concern. Most drugs are not completely metabolized by humans and animals, and a significant percentage of administered antibiotics enter anaerobic bioreactors with the processed feedstock, including agricultural residues, animal manure, and sewage sludge [1]. The concentrations of antimicrobial residues in raw substrates and digestates vary significantly and range from nanograms to micrograms per kg or mL. Although antibiotics can be degraded during AD, drug removal rates vary considerably due to differences in treatment parameters, such as temperature, organic load rate, or hydraulic retention time [87,88]. Moreover, antibiotic degradation rates during AD are influenced by the chemical structure of the drug, as well as the type and characteristic parameters of the feedstock. The reported rates of drug degradation differ between studies [24,81,89]. The efficiency of drug removal is determined by process conditions and the type of processed raw material, and it has been estimated at 7–98% for tetracyclines, 36–95% for MLS antibiotics, and 20–83% for fluoroquinolones [22,37,90]. Antibiotics persist in AD by-products, which can potentially lead to the emergence, persistence, or accumulation of these micropollutants in the environment when digestate is applied to soil [87,91]. The persistence and transformation of antibiotic residues in soil are influenced by numerous processes. Various antibiotic degradation rates in soil have been reported in the literature [87,92]. Digestate can be applied as a soil conditioner to promote sustainable biogas production and the circular economy, which is why the fate of antibiotics in substrates processed by AD should be analyzed in greater detail. The risks associated with drug residues in AD by-products have not been sufficiently investigated to date, which highlights the need for further research [93]. Therefore, the types and concentrations of antimicrobials identified in digestates derived from various feedstocks are presented in

Table 3. Types and concentrations of antimicrobial substances detected in digestates derived from the AD of various substrates.

Antimicrobial Class and Substance		Concentration of Antimicrobial Substance: Average Value or Minimum−Maximum Range		Substrate	Country	References
β-lactams	Amoxicillin	460–960	μg kg−1 DM	Food waste	Norway	[94]
	penicillin G	510	μg kg−1 DM	Cattle manure
Fluoroquinolones	Ciprofloxacin	2100	μg kg−1		Italy	[12]
		776,000	μg kg−1 DM	Mixed animal substrates	Sweden, Finland, and Germany	[95]
		1620	μg kg−1 DM		Germany	[96]
		63,100	μg kg−1 DM	Mixed pig slurry and sewage sludge	Sweden, Finland, and Germany	[95]
		121,000	μg kg−1 DM	Poultry manure
		430	μg kg−1 DM	Sewage sludge	Norway	[94]
		16.05–45.26	μg kg−1		China	[97]
		1068	μg L−1	Pig manure		[98]
		57.3–122.4	μg kg−1			[99]
	Danofloxacin	963	μg L−1			[98]
		970	μg kg−1 DM	Mixed animal substrates	Germany	[96]
	Enrofloxacin	147,000	μg kg−1 DM	Cattle manure	Sweden, Finland, and Germany	[95]
		82,100	μg kg−1 DM	Mixed animal substrates
		Max 1090	μg kg−1 DM		Germany	[96]
		50	μg kg−1	Mixed plants and animal manure	Poland	[100]
		47,700	μg kg−1 DM	Mix pig slurry and sewage sludge	Sweden, Finland, and Germany	[95]
		850,000	μg kg−1 DM	Poultry manure
		6.28–20.28	μg kg−1	Sewage sludge	China	[97]
		147.1–387.9	μg kg−1		Poland	[11]
		200–44,200	μg kg−1 DM	Pig manure	Sweden, Finland, and Germany	[95,101]
		98	μg L−1		China	[98]
		58.7–61	μg kg−1			[99]
	Fleroxacin	127	μg L−1			[98]
	Lomefloxacin	409	μg L−1
	Norfloxacin	132.54–444.27	μg kg−1			[97]
		55.9–82.4	μg kg−1			[99]
	Ofloxacin	628.89–1751.08	μg kg−1			[97]
		49.9–60.2	μg kg−1			[99]
		184	μg L−1			[98]
	Orbifloxacin	33	μg L−1
	Sarafloxacin	445	μg L−1
	Sparfloxacin	140	μg L−1
MLS	Clarithromycin	353.28–369.84	μg L−1	Pig manure	Denmark	[102]
		0.056–0.208	μg L−1 (LF)	Sewage sludge	Poland	[11]
		62.8 70.7	μg kg−1 (SF)
	Clindamycin	0.08–0.268	μg L−1 (LF)
	Erythromycin	7.8–624	μg L−1	Pig manure	Denmark	[102]
Pleuromutilins	Tiamulin	148	μg kg−1	Mixed plants and animal manure	Poland	[100]
Polyethermonocarboxylic acids	Monensin	220–720	μg kg−1	Cattle manure	USA	[103]
Sulfonamides	Sulfadiazine	233,000	μg kg−1 DM		Sweden, Finland, and Germany	[95]
		n.d.–0.106	μg L−1 (LF)	Mixed cattle manure and maize silage	Poland	[11]
		105,000	μg kg−1 DM	Mixed animal substrates	Sweden, Finland, and Germany	[95]
		26,300	μg kg−1 DM	Mixed pig slurry and sewage sludge
		239,000	μg kg−1 DM	Poultry manure	Sweden, Finland, and Germany	[95,96]
		0.098–0.136	μg L−1 (LF)	Sewage sludge	Poland	[11]
		12,400	μg kg−1 DM		Sweden, Finland, and Germany	[95]
		142,000	μg kg−1 DM	Pig manure
		618	μg L−1		Denmark	[102]
		n.d.–4.9	μg kg−1		China	[99]
	Sulfadoxine	n.d.	μg kg−1 DM		Germany	[101]
		140	μg kg−1 DM	Food waste	Norway	[94]
	Sulfamethoxazole	70	μg kg−1	Cattle manure	Italy	[12]
		n.d.–7.4	μg kg−1	Pig manure	China	[99]
	Sulfamethazine	1400–200,900	μg kg−1 DM	Mixed animal substrates	Germany	[104]
		1.6	μg kg−1 DM	Mixed cattle and pig manure	Norway	[94]
		0.38	μg kg−1 DM	Mixed food waste and sewage sludge
		66,100	μg kg−1 DM	Poultry manure	Sweden, Finland, Germany	[95]
		0.08	μg kg−1 DM	Sewage sludge	Norway	[94]
		176–3359.9	μg kg−1	Pig manure	China	[99]
Tetracyclines	Chlortetracycline	240–340	μg kg−1	Cattle manure	USA	[105]
		3500–36,500	μg kg−1 DM	Mixed animal substrates	Germany	[104]
		1300–10,100	μg kg−1 DM	Pig manure		[101]
		n.d.	μg kg−1 TS		China	[25]
		11,600	μg L−1 DM		USA	[106]
		1000–6000	μg L−1			[103]
		36 ± 2	μg L−1		China	[98]
		21,010.6–39,751.4	μg kg−1			[99]
	Doxycycline	0.854–1.555	μg L−1 (LF)	Mixed cattle manure and maize silage	Poland	[11]
		396.7–1282.5	μg kg−1 (SF)
		n.d.–3900	μg kg−1 DM	Pig manure	Germany	[101,107]
		360	μg L−1		The Netherlands	[108]
		0.62 218.1–800.2	μg L−1 (LF) μg kg−1 (SF)	Sewage sludge	Poland	[11]
	Epichlortetracycline	99–170	μg kg−1 TF	Cattle manure	USA	[105]
	Epitetracycline	83–102	µg kg−1 TF
	Oxytetracycline	196,000	μg kg−1 DM		Sweden, Finland, and Germany	[95]
		21–26	μg kg−1 TF		USA	[105]
		196,000	μg kg−1 DM	Mixed animal substrates	Sweden, Finland, and Germany	[95]
		346,000	μg kg−1 DM	Poultry manure
		1,256,000	μg kg−1 DM	Pig manure
		n.d.	μg kg−1 TS		China	[25]
		34,911.8–84,687.9	μg kg−1			[99]
		38.5	μg L−1		The Netherlands	[108]
	Tetracycline	290–450	μg kg−1 TF	Cattle manure	USA	[105]
		n.d.–17,030	μg kg−1 DM	Mixed animal substrates	Germany	[96,104]
		193,000	μg kg−1 DM	Mixed animal substrates	Sweden, Finland, and Germany	[95]
		438,000	μg kg−1 DM	Mixed pig slurry and sewage sludge
		16,399,000	μg kg−1 DM	Poultry manure
		0.27–0.36 464.8–1164.4	μg L−1 (LF) μg kg−1 (SF)	Sewage sludge	Poland	[11]
		885,000	μg kg−1 DM		Sweden, Finland, and Germany	[95]
		n.d.–6400	μg kg−1 DM	Pig manure	Germany	[101]
		1209.1–1769.7	μg kg−1		China	[99]
		42.14	μg L−1		The Netherlands	[108]

DM—dry matter; TF—total fraction; TSs—total solids; LF—liquid fraction; SF—solid fraction; n.d.—not detected.

.

The role of digestate as a source of antibiotics in soil ecosystems needs to be clarified. As indicated in Table 3, the concentrations of antimicrobial substances differ widely in various digestates. Antibiotic levels in digestates are determined mainly by the quality of the substrate. Sewage sludge and animal manure are more likely to be contaminated with drugs than agricultural residues. It has been reported that sulfonamide and fluoroquinolone antimicrobial agents are most prevalent in animal manure and sewage sludge [12,49,109]. Research has also shown that tetracycline and sulfonamide drugs can persist after the AD process [91]. As shown in Table 3, fluoroquinolone, sulfonamide, and tetracycline antibiotics have been most frequently detected in various AD by-products. However, the influence of AD on antibiotic concentrations in the substrates processed in anaerobic bioreactors remains unclear [12]. According to the literature, antimicrobial agents are degraded during anaerobic treatment, but this process has not been specifically designed to remove drugs [110]. As mentioned previously, the efficiency of antibiotic degradation depends on many factors. In some studies, antibiotics were not effectively removed during the AD process [104,111]. In turn, other authors have reported very high drug removal rates during AD and, consequently, very low drug concentrations in the obtained digestates [112].

The efficiency of drug removal and the concentrations of drug residues in digestates are also highly dependent on the chemical structure and specificity of antibiotics. Some antimicrobial substances are removed effectively, but their removal rates are very low (such as chlortetracycline and oxytetracycline that are characterized by high adsorption rates due to the presence of chlorine atoms and hydroxyl groups). Fluoroquinolones have a lower removal rate due to the presence of a fluorine atom [110]. In turn, β-lactams contain highly unstable β-lactam rings that are degraded by microbial β-lactamases during AD [113]. Moreover, antibiotic removal is considerably influenced by AD conditions. It has been reported that drug removal efficiency is strictly dependent on the total content of solids in the feedstock and the temperature of the anaerobic process [110]. Low drug removal rates are responsible for high concentrations of drug residues in AD by-products. The levels of ciprofloxacin, enrofloxacin, sulfadiazine, and oxytetracycline residues were particularly high in various digestates (Table 3). Due to their high stability and high adsorption potential, these antibiotics can persist in soil for long periods of time, thus posing a serious threat to the environment.

The lack of legal regulations concerning safe concentrations of drugs introduced into the soil with digestate gives serious cause for concern. Although the potential of digestate as an organic fertilizer has been studied to determine contamination with pathogens and HMs, as well as nutrient levels, the risks associated with antibiotic residues in digestate have been overlooked [21]. In Europe, the use of organic fertilizers (including digestate) is promoted by the EU Fertilizing Products Regulation (EU 2019/1009) which establishes the threshold values for total nitrogen, phosphorous, organic carbon, pathogens, and HMs that affect fertilizer quality. However, antibiotic residue levels have not been regulated to date [12]. Legal regulations concerning the permissible levels of antibiotics in digestates intended for soil fertilization should be urgently introduced. To the best of our knowledge, the presence of some antimicrobials should be banned, whereas the concentrations of other antibiotics should be restricted in fertilizer products (by defining limiting concentrations, as in the case of HMs). These measures are urgently needed to prevent the accumulation of pharmaceuticals in the environment and mitigate their negative effects on public health. At the beginning of 2024, the World Health Organization (WHO) published an updated list of medically important antimicrobials (MIAs) [114]. These drugs have been listed based on their importance in medicine, risk of antimicrobial resistance, and potential implications for public health resulting from inappropriate use, particularly in livestock farming. We believe that digestates containing third- and fourth-generation cephalosporins, quinolones, and polymyxins, i.e., drugs considered “the highest priority critically important antimicrobials”, should be banned from agricultural use. In addition, maximum permissible levels should be introduced for antimicrobials classified as “critically important” and “highly important”, especially aminoglycosides, macrolides, lincosamides, streptogramins, penicillins, pleuromutilins, sulfonamides, and tetracyclines, which were often detected in various digestates (Table 3).

3.3. Ecological Risks Associated with HM and Antibiotic Contamination

All HMs and some antimicrobial agents are considered persistent pollutants in agricultural ecosystems because long-term accumulation of these substances in the soil not only exerts highly toxic effects on various organisms and plants, but also contributes to the spread of microbial resistance [110]. When digestates contaminated with HMs or/and antibiotics are used as agricultural fertilizers, these substances pollute the environment and exert selective pressure on soil microorganisms [21,88]. Digestates and soil are colonized by diverse microorganisms, where bacteria harboring antibiotic resistance genes (ARGs) and heavy metal resistance genes (MRGs) pose a particular threat. Due to the widespread use of antibiotics in recent decades, ARGs have been classified as a new source of pollution that poses a threat to public health and safety [115]. High consumption of antibiotics promotes the emergence and persistence of antimicrobial resistance. Drugs exert selective pressure on ARGs [116]. In addition, HMs cannot be biodegraded; therefore, the selection pressure exerted on microorganisms in their presence is long-standing. It has been noted that high concentrations of HMs induce metal tolerance in communities of soil-dwelling bacteria [117]. Moreover, HMs also contribute to the spread of antimicrobial resistance. Research has shown that HM contamination can also promote the dissemination of ARGs among bacteria. At the molecular level, these phenomena can be interpreted as co-selection [118]. The presence of HMs in ecosystems can accelerate the development and propagation of ARGs [119], and the emergence of some ARGs may be directly associated with the presence of HMs in the environment [120]. There is scientific evidence to suggest that the release of HMs and antimicrobial substances into the environment is correlated with the presence of MRGs and ARGs. The co-selection of ARGs and MRGs has been reported in various environments [120,121,122,123]. In the context of digestate contamination, the impact of HMs and drugs on the soil environment is most meaningful because anaerobically stabilized organic matter is used primarily as a soil conditioner and fertilizer [124,125,126,127]. In one of the reviewed studies, the abundance of ARGs in the soil increased after exposure to HMs, and the observed increase was proportional to the rise in HM concentrations [128]. In another study, the spread of ARGs increased after HMs was released into antibiotic-contaminated soils [129]. However, HMs and antimicrobials may exert synergistic or antagonistic effects in the soil environment, which affects the correlation between ARG abundance and antibiotic levels [130]. Digestate-based fertilizers can release HMs and antimicrobial substances, as well as promote the spread of MRGs and ARGs in cultivated fields. These micropollutants may be transferred from the soil to groundwater and crops, ultimately reaching humans and animals [131].

In the present study, the specific environmental risks associated with HMs and drugs were analyzed to determine the potential hazards resulting from the introduction of digestates into the soil. The levels of HM contamination were estimated by calculating the geo-accumulation index (I_geo) and the ecological risk factor (E_r) which indicate the extent to which the analyzed digestates induce changes in HM concentrations in soil and exert toxic effects on the ecosystem (Figure 3A,B, respectively). The values of I_geo and E_r calculated for Cd, Cu, Hg, Mo, and Ni were indicative of high pollution levels and a high ecological risk. Approximately 25% to 33% of the analyzed digestates were extremely contaminated with Cd and Hg (I_geo > 5), and approximately 45% to 50% of the digestates were characterized as posing a very high ecological risk (E_r > 320). Contamination with Mo was analyzed in only one publication, and the examined digestate was characterized by the highest level of pollution which exerted a potential risk for the ecosystem. None of the analyzed digestates were contaminated with Co, Fe, or Mn (I_geo < 1), and the concentrations of these HMs were indicative of low ecological risk (E_r < 40). In turn, the RQ method was used to assess the extent to which the presence of drug residues in various digestates can affect selection for antibiotic resistance in the environment (RQ_mic; Figure 3C) and ecotoxicity (RQ_env; Figure 3D) of fertilizer-amended soil. In most digestates (63–100%), the RQ_mic values for β-lactams (amoxicillin), fluoroquinolones (ciprofloxacin, enrofloxacin, and sparfloxacin), and tetracyclines (oxytetracycline) were considerably higher than 1, which points to a high risk of selection for antimicrobial resistance in soil. Drugs representing the same antibiotic classes (β-lactams—amoxicillin; fluoroquinolones—ciprofloxacin and enrofloxacin; tetracyclines—oxytetracycline and tetracycline) posed a high risk for soil-dwelling organisms (RQ_env > 1 in 33–100% of digestate samples). In turn, all RQ_mic values for clindamycin, erythromycin, and sulfamethoxazole (in the range of 0.0001–0.91), as well as RQ_env values for clindamycin, norfloxacin, and sulfamethoxazole (0.0006–0.37), were indicative of no risk or low risk for soils fertilized with various digestates.

The environmental risk assessment demonstrated that most digestates containing HMs and antibiotics could pose a high risk to the ecosystem and promote the spread of antimicrobial resistance if introduced into the soil. Moreover, the antibiotic classes considered critically important (quinolones) and highly important (β-lactams and tetracyclines) in the WHO MIA List tend to be present in high concentrations in various digestates and may exceed the alert limit as regards the risk of selection for drug resistance. The above reiterates the need for the establishment of legal limits concerning the maximum admissible concentrations of individual drugs in anaerobically stabilized organic matter that can be released into the soil environment. Considering the alarming levels of anthropogenic pollution, the fact that soil contamination with HMs and antimicrobials can exert long-term effects and promote widespread co-selection pressure for microbial resistance is of particular concern. Therefore, the close links between ARGs, MRGs and their co-transfer, especially in the presence of HMs and antibiotics, pose the greatest challenge for research in the field of environmental microbiology. The role of digestate in the environmental transfer of HMs and antibiotics is extremely important; therefore, the residual levels of these micropollutants in AD by-products should be systematically monitored and analyzed.

4. Conclusions and Future Perspectives

This review article analyzes the global scope of digestate contamination with HMs and antibiotics. Based on the results of the analysis, it can be concluded that the AD process and the resulting digestates directly contribute to the presence of HMs and antibiotics in the environment. The absence of effective methods for removing these micropollutants during anaerobic treatment promotes their further transfer into the ecosystem. Although digestates generally meet quality standards, their application in farming could pose a threat to the soil and water environment and, subsequently, to public health. This review demonstrated that HM and antibiotic residues in digestates pose a risk for soil-dwelling organisms and contribute to selection for microbial resistance in soil ecosystems. The presence of antimicrobial substances in digestates intended for agricultural use should be urgently addressed by legal regulations. We believe that a better understanding of the role of digestate as a source of anthropogenic micropollutants and the risks associated with its introduction into the soil environment will encourage a scientific debate and lead to the implementation of dedicated legislative initiatives.