3.1. Carbon mineralization patterns in native and residue-amended soils
The cumulative amounts of CO2-C released from the native and residue-amended soils are shown in Fig. 1. The CO2-C evolution from the native soil was relatively low during the incubation period. For instance, the total cumulative mineralized C (Cmin) in native soil reached 298.5 mg CO2-C kg− 1, indicating that on average only 4.66% of the native soil organic carbon was mineralized under the optimal soil moisture and temperature conditions applied in the incubation experiment. Moreover, the initial C mineralization rate (the CO2-C release rate on the first day of the incubation period) was 44.2 mg CO2-C kg− 1 day− 1 in the native soil, which reached the lowest value of 1.66 mg CO2-C kg− 1 day− 1 at the end of the incubation period.
Addition of plant residues to the soil significantly (P < 0.05) increased the rate and quantity of CO2-C evolution from the soil, as shown in Fig. 1. The results revealed a rapid phase of C mineralization occurred in the first weeks of incubation presumably performed by the copiotrophic microbes followed by a slow phase of C mineralization mainly accomplished by the oligotrophic microbial community. The copiotrophic (zymogemous) microorganisms tend to proliferate when a readily-available organic matter is incorporated into the soil, while the oligotrophic (autochthonous) microorganisms are principally responsible for the decomposition of more recalcitrant organic substances in situations where available C depletion occurs through competition among microorganisms (Dilly, 2005; Yarwood, 2021). At any given time of the incubation period, the C mineralization was significantly (P < 0.05) higher in the residue-amended soils compared to that of the control (unamended) soil (Fig. 1).
The amount of C mineralized was significantly (P < 0.05) greater in the soil amended with AH than that amended with WS, during the whole incubation experiment (Fig. 1), which could be related to a larger proportion of the readily degradable C in the AH residues (Pascault et al., 2010). The AH and WS residues significantly differed in their C/N ratios (Table 1), which were predicted to influence C mineralization rates. In a similar work, the higher amount of CO2-C evolved from an AH-amended soil compared to a WS-amended soil was attributed to the greater biodegradability of AH residues caused by the lower C/N and lignin:N ratios of the AH residues (Nourbakhsh and Sheikh-Hosseini, 2006). The Cmin value was calculated as 5032 and 4697 mg CO2-C kg− 1 for the AH- and WS-amended soils, respectively (Fig. 1). Assuming a negligible priming effect, the net mineralized C was 47.3% and 44.0% of the originally added C by AH and WS residues, respectively.
The initial C mineralization rate was also higher in the soil amended with AH compared to the soil amended with WS, as follows from the initial slopes of the C mineralization curves in Fig. 1. The initial C mineralization rate in soil amended with AH was 774.0 mg kg− 1 d− 1 which was significantly (P < 0.05) higher than the value of 300.0 mg kg− 1 d− 1 recorded for WS-treated soil. The C mineralization rate values decreased with time and finally reached 49.14 and 21.85 mg kg− 1 d− 1, in AH- and WS-amended soils, respectively, at the end of the incubation period, indicating a decrease of 93.7% and 92.7% compared to the first day.
3.2. Effects of ZnO-NPs on C mineralization pattern
The effects of various concentrations of ZnO-NPs on the daily and cumulative C mineralization pattern are shown in Figs. 2 and 3, respectively. Generally, all treated soils showed a fast initial CO2-C evolution because of the rapid depletion of the readily-mineralizable substances, followed by a slower release throughout the remaining period of incubation. The appearance of secondary maxima (Fig. 1) may be due to the variety of compounds present in the residues and their dissimilar degrees of degradability (Bernal et al., 1998; Martín et al., 2012).
Carbon mineralization trends over time were similar for the uncontaminated soil samples and soils contaminated with different ZnO-NPs concentrations. However, ZnO-NPs posed some changes in the C mineralization pattern during the incubation period. As a result, the differences between NPs treated and untreated soils generally widened with the time of incubation (Fig. 2).
The total cumulative mineralized C (Cmin) value was significantly (P < 0.05) decreased in the AH-amended soil at the lower ZnO-NPs concentrations (100 and 200 mg kg− 1 soil) as compared to the control, while at greater concentrations of ZnO-NPs no significant differences were observed in the Cmin values among the contaminated and control soil samples (Fig. 4A). In the WS-amended soil samples, the Cmin value was also significantly (P < 0.05) decreased at all ZnO-NPs concentrations, except the lowest concentration of 100 mg kg− 1 soil (Fig. 4. B). The reductions observed in Cmin value can be related to a lower decomposer community in the presence of ZnO-NPs as previous studies have indicated that the decomposer microbial community is significantly vulnerable to ZnO-NPs (Collins et al., 2012; Ge et al., 2011; Shah et al., 2022). This finding is in agreement with that of Shemawar et al. (2021), indicating that leaf litter C mineralization was decreased after ZnO-NPs addition to mesocosms containing sandy soil. García-Gómez et al. (2015) also showed a substantial decline in the C mineralization of glucose in soils treated with ZnO-NPs. They attributed the decrease in glucose-C mineralization to the eradication of soil organisms sensitive to ZnO-NPs.
On the other hand, the unchanged microbial respiration in the presence of high ZnO-NPs concentration in AH-amended soil (Fig. 4A) could have resulted from the higher energy requirements for the maintenance of an attenuated microbial population under ZnO-NPs stress. Tobor-Kapłon et al. (2005) declared that under stress conditions, more resistant microorganisms respond by intensified respiration, while more sensitive microorganisms are recognized by decreased respiration. Such dissimilar results obtained for AH- and WS-amended soil samples can also be attributed to the different biochemical properties of available substrates in plant residues that are mineralized (Giller et al., 1998).
3.3. Effects of ZnO-NPs on carbon mineralization kinetics
The presence of NPs in the soil can significantly influence the ability of the soil microorganisms to mineralize organic substances, and therefore a measurement of soil respiration kinetics is a valuable method to assess the impacts of NPs on soil microbial activity (Shemawar et al., 2021; Xin et al., 2020). In this study, the time-dependent organic C mineralization data were fitted by the first-order, double first-order, first-order E, and special models. The statistical parameters obtained through the non-linear regression analysis of the kinetic data derived for the soils amended with AH and WS are presented in Tables 3 and 4, respectively.
Table 3
Adjusted coefficients of determination (Adj. R2), root mean squared errors (RMSE), corrected Akaike information criteria (AICc), and AICc differences (∆AICc) for the models describing C mineralization kinetics of alfalfa hay in soil samples treated with different concentrations of ZnO-NPs.
Concentration of ZnO-NPs
|
First-order
|
First-order E
|
Special model
|
Double first order
|
Adj. R2
|
RMSE
|
AICc
|
∆AICc
|
Adj. R2
|
RMSE
|
AICc
|
∆AICc
|
Adj. R2
|
RMSE
|
AICc
|
∆AICc
|
Adj. R2
|
RMSE
|
AICc
|
∆AICc
|
0
|
0.950
|
274.6
|
544.6
|
72.10
|
0.970
|
209.8
|
521.1
|
48.60
|
0.986
|
144.2
|
485.1
|
12.60
|
0.989
|
123.2
|
472.5
|
00.00
|
100
|
0.944
|
256.5
|
538.0
|
43.90
|
0.961
|
209.9
|
521.2
|
27.10
|
0.978
|
164.3
|
497.7
|
03.60
|
0.980
|
154.2
|
494.1
|
00.00
|
200
|
0.952
|
241.2
|
532.1
|
55.00
|
0.972
|
186.2
|
509.7
|
32.60
|
0.982
|
149.4
|
488.6
|
11.50
|
0.986
|
129.2
|
477.1
|
00.00
|
400
|
0.960
|
240.6
|
531.9
|
91.20
|
0.978
|
179.5
|
506.2
|
65.50
|
0.992
|
110.5
|
459.6
|
18.90
|
0.994
|
88.46
|
440.7
|
00.00
|
600
|
0.967
|
226.6
|
526.2
|
160.5
|
0.979
|
181.8
|
507.4
|
105.7
|
0.995
|
86.14
|
435.7
|
34.00
|
0.998
|
58.91
|
401.7
|
00.00
|
800
|
0.967
|
223.7
|
524.9
|
132.8
|
0.981
|
172.9
|
502.6
|
110.5
|
0.995
|
83.56
|
432.8
|
40.70
|
0.998
|
53.31
|
392.1
|
00.00
|
1000
|
0.960
|
251.1
|
536.0
|
52.80
|
0.971
|
215.4
|
523.7
|
40.50
|
0.986
|
151.2
|
489.7
|
06.50
|
0.987
|
137.6
|
483.2
|
00.00
|
Table 4
Adjusted coefficients of determination (Adj. R2), root mean squared errors (RMSE), corrected Akaike information criteria (AICc), and AICc differences (∆AICc) for the models describing C mineralization kinetics of wheat straw in soil samples treated with different concentrations of ZnO-NPs.
Concentration of ZnO-NPs
|
First-order
|
First-order E
|
Special model
|
Double first order
|
Adj. R2
|
RMSE
|
AICc
|
∆AICc
|
Adj. R2
|
RMSE
|
AICc
|
∆AICc
|
Adj. R2
|
RMSE
|
AICc
|
∆AICc
|
Adj. R2
|
RMSE
|
AICc
|
∆AICc
|
0
|
0.992
|
120.1
|
465.2
|
21.50
|
0.995
|
95.24
|
445.3
|
01.60
|
0.994
|
106.9
|
456.4
|
12.70
|
0.996
|
91.23
|
443.7
|
00.00
|
100
|
0.979
|
195.0
|
511.7
|
01.70
|
0.981
|
186.7
|
510.0
|
00.00
|
0.979
|
193.0
|
513.1
|
03.10
|
0.981
|
186.7
|
512.4
|
02.40
|
200
|
0.969
|
197.8
|
513.1
|
01.60
|
0.970
|
189.9
|
511.6
|
0.01
|
0.972
|
189.7
|
511.5
|
00.00
|
0.971
|
186.3
|
512.3
|
0.80
|
400
|
0.978
|
167.6
|
497.2
|
07.00
|
0.981
|
154.4
|
491.7
|
1.50
|
0.982
|
152.0
|
490.2
|
00.00
|
0.981
|
148.2
|
490.3
|
0.10
|
600
|
0.992
|
100.3
|
447.9
|
39.70
|
0.996
|
71.33
|
417.6
|
09.40
|
0.996
|
72.05
|
418.5
|
10.30
|
0.997
|
63.00
|
408.2
|
00.00
|
800
|
0.984
|
130.7
|
473.3
|
30.40
|
0.990
|
101.9
|
451.8
|
08.90
|
0.991
|
98.56
|
448.6
|
05.70
|
0.992
|
90.50
|
442.9
|
00.00
|
1000
|
0.978
|
163.2
|
494.7
|
24.50
|
0.985
|
134.1
|
478.2
|
08.00
|
0.986
|
128.0
|
473.7
|
03.50
|
0.987
|
120.2
|
470.2
|
00.00
|
Based on the adjusted R2 values, all tested kinetic models significantly fitted the time-dependent C mineralization data. However, based on the AICc values, we selected the double first-order model to further describe the C mineralization kinetics. Referring to Table 3, the double first-order model was strongly favored (ΔAICc = 0) over the others to account for C mineralization kinetics in the AH-amended soil at all ZnO-NPs concentrations. The other three models showed ΔAICc values mainly greater than 10, indicating that they have substantially less support compared to the best model. The double first-order model also showed the lowest AICc score and therefore was the best model to explain the C mineralization kinetics, in WS-amended soil treated with all ZnO-NPs concentrations, except for 100, 200, and 400 mg kg− 1 (Table 4). Nevertheless, for these treatments the double first-order model was also very competitive, representing the ΔAICc values of 2.40, 0.80, and 0.10, respectively. Based on these results, and for the sake of consistency, the double first-order model was selected as the most appropriate model for describing the C mineralization kinetics in this study. The double first-order model has also been successfully used in other studies to describe the decomposition of organic substances in soil (Barral et al., 2009; Moreno-Cornejo et al., 2015; Shi et al., 2019).
The adjustable parameters obtained from the fitted double first-order model are useful to estimate the sizes and mineralization rates of the labile and recalcitrant organic C pools in the residue-amended soils. The kinetic parameters derived from the double first-order model are represented in Table 5. According to the results, the size of the labile C that mineralized during the first rapid stage (C1) was 2411 mg kg− 1 for the uncontaminated AH-amended soil. The C1 value dropped slightly when 100 and 200 mg ZnO-NPs kg− 1 were added to the AH-amended soil, and then sharply increased with further rising ZnO-NPs concentration (Table 5). The maximum C1 value in AH-amended soil was 2806 mg kg− 1 observed at the highest ZnO-NPs application level (1000 mg kg− 1). For the WS-amended soils, the C1 value was 278.0 mg kg− 1 in the uncontaminated soil sample and constantly increased with increasing the degree of soil contamination with ZnO-NPs. The highest C1 value in the WS-amended soil sample was found in the presence of 1000 mg ZnO-NPs kg− 1 soil (Table 5). Moreover, the C1 value was significantly correlated to the rate of applied ZnO-NPs in both AH-amended (r = 0.87, P < 0.05) and WS-amended (r = 0.89, P < 0.05) soils, suggesting an intensification of respiratory C mineralization by soil copiotrophic populations to acquire the higher energy required for their existence as ZnO-NPs stress increased (Anderson and Domsch, 2010).
Table 5
Kinetic parameters of the double first-order model describing C mineralization of alfalfa hay and wheat straw in soil samples treated with different concentrations of ZnO-NPs.
Model parameter
|
Concentration of ZnO-NPs
|
0
|
100
|
200
|
400
|
600
|
800
|
1000
|
|
Alfalfa hay
|
C1 (mg kg− 1)
|
2411
|
2381
|
2216
|
2469
|
2784
|
2736
|
2806
|
k1 (day− 1)
|
0.3268
|
0.3065
|
0.3440
|
0.3064
|
0.2908
|
0.2941
|
0.2863
|
C2 (mg kg− 1)
|
2997
|
2449
|
2520
|
2803
|
2652
|
2647
|
2577
|
k2 (day− 1)
|
0.0357
|
0.0347
|
0.0460
|
0.0362
|
0.0349
|
0.0362
|
0.0360
|
Ct (C1 + C2, mg kg− 1)
|
5408
|
4830
|
4736
|
5272
|
5436
|
5383
|
5383
|
|
Wheat straw
|
C1 (mg kg− 1)
|
278.0
|
152.9
|
371.5
|
532.5
|
420.1
|
559.2
|
688.7
|
k1 (day− 1)
|
0.5124
|
1.403
|
0.2909
|
0.2418
|
0.2772
|
0.2658
|
0.2683
|
C2 (mg kg− 1)
|
5363a
|
4956ab
|
4054bc
|
3991c
|
4488abc
|
3900c
|
3908c
|
k2 (day− 1)
|
0.0302
|
0.0368
|
0.0329
|
0.0313
|
0.0267
|
0.0272
|
0.0289
|
Ct (C1 + C2, mg kg− 1)
|
5641
|
5109
|
4425
|
4523
|
4908
|
4459
|
4597
|
The size of the recalcitrant C pool (C2), which was slowly mineralized during the second phase of the incubation experiment, varied between 2449 to 2997 mg kg− 1 in the AH-amended soils and between 3900 to 5363 mg kg− 1 in the WS-amended soils. The C2 constant showed its maximum value in uncontaminated samples in both AH- and WS-amended soil samples and significantly decreased in the presence of ZnO-NPs (Table 5). The C2 value showed a significant inverse correlation with the rate of applied ZnO-NPs in the WS-amended soil sample (r = − 0.73, P < 0.05), while a similar relationship was not statistically significant for the AH-amended soil. The lower values of C2 in soils containing ZnO-NPs reflect that the activity of soil oligotrophic populations, mineralizing the more recalcitrant fractions of plant residues, deteriorated in the presence of ZnO-NPs, as this community is not likely able to boost its respiratory activity for obtaining more energy to resist the ZnO-NPs toxicity (Yarwood, 2021).
The total potentially mineralizable C (Ct) values, obtained from the summation of C1 and C2 constants, ranged from 4736 to 6436 mg kg− 1 for soil amended with AH and from 4425 to 5641 mg kg− 1 for soil amended with WS (Table 5). In AH-amended soil, the labile C fraction (C1) consisted of 48.8% of the Ct, while in WS it was only 9.2% of the total mineralizable Ct. This implies that a high percentage (90.8%) of the C mineralization of WS residue took place during the second slow phase since WS has a much larger amount of recalcitrant compounds than AH (Nourbakhsh and Sheikh-Hosseini, 2006; Pascault et al., 2010). In addition, the values of the potential mineralizable C (Ct) were always higher than the measured cumulative mineralized C (Cmin) obtained in the incubation period, confirming that more decomposable organic C remained in the residue-amended soil after 58 days.
The impacts of ZnO-NPs on the C mineralization rate constants of the fast pool (k1) and the slow pool (k2) are represented in Table 5. In general, the addition of ZnO-NPs decreased the k1 value in both AH- and WS-amended soil samples. This indicates a slower mineralization rate (i.e., longer residence time) of labile C substances in contaminated soil during the initial phase of mineralization, probably due to the toxic impacts of ZnO-NPs on soil zymogenous microbial community. However, no specific trend was noted in the k2 value by increasing the concentration of ZnO-NPs. For instance, the k2 value remained nearly unchanged in AH-amended soil, in the presence of all ZnO-NPs concentrations, except 200 mg kg− 1. In the WS-amended soil samples, the k2 value first increased by treatment with 100 to 400 mg ZnO-NPs kg− 1 soil and then decreased with the application of the higher ZnO-NPs concentrations (Table 5), implying that the rate of recalcitrant C mineralization by oligotrophic microflora significantly reduced only by high concentrations of ZnO-NPs in WS-amended soil.
3.4. Effects of ZnO-NPs on soil microbial biomass
The results revealed significant differences in microbial biomass carbon (MBC) among soils affected by ZnO-NPs contamination and the control (Fig. 5). In the AH-amended uncontaminated soil sample, MBC was 333.3 mg kg− 1, which significantly (P < 0.05) decreased by all concentrations of ZnO-NPs greater than 100 mg kg− 1. The lowest MBC was 156.6 mg kg− 1, recorded in the soil containing 1000 mg ZnO-NPs kg− 1. In the WS-amended soil, the MBC value was 500.0 mg kg− 1 in the control treatment which decreased significantly (P < 0.05) in the presence of 400 mg kg− 1 and greater concentrations of Zn-NPs (Fig. 5). The decrease of MBC was by 8.34 to 53.3% in AH-amended soil and by 4.44 to 54.4% in WS-amended soil, depending on the concentration of ZnO-NPs added. This finding supports our hypothesis that ZnO-NPs could adversely affect soil microbial biomass. The results are consistent with other studies in which ZnO-NPs had reduced soil microbial abundance. Rashid et al. (2017), for example, reported that the application of 1000 mg ZnO-NPs kg− 1 in a litter-amended soil significantly reduced the heterotrophic bacterial and fungal populations and MBC compared to control soil. Shemawar et al. (2021) also observed a significant MBC reduction in a calcareous soil spiked with 100 and 1000 mg kg− 1 ZnO-NPs in 24 days. The reduced soil microbial biomass by ZnO-NPs could be attributed to the direct killing of microorganisms through ROS production, protein misfolding, mitochondrial damage, DNA fragmentation, and membrane disruption (Raffi and Husen, 2019; Xin et al., 2020). NPs may also indirectly affect soil microbial biomass by altering the soil environment such as water properties that are vital for microbial survival and proliferation (Ge et al., 2013).
3.5. Effects of ZnO-NPs on soil basal respiration
The basal respiration (BR) value in AH-amended soil decreased by all ZnO-NPs concentrations as compared to that of the control, however, the reduction was statistically significant only at 200 mg ZnO-NPs kg− 1 soil (Fig. 5). In WS-amended soil, BR was significantly (P < 0.05) decreased in the presence of all ZnO-NPs concentrations, except 100 and 600 mg kg− 1 soil (Fig. 5). Depending on the concentration of ZnO-NPs applied, the BR value was decreased by 7.33–35.7% in AH-amended soil and by 5.13–34.5% in WS-amended soils, compared to those of the corresponding uncontaminated soil samples. Based on these findings, it can be concluded that the BR value of the soils generally decreased with increasing ZnO-NPs concentration; however, the trend of this reduction was less severe as compared to that of the MBC (Fig. 5). Shifting in the microbial community composition has been suggested as a possible reason as to why changes in MBC induced by NPs did not change soil BR substantially (Ge et al., 2011). For instance, Zhou et al. (2020) showed that after a 70-day exposure to 2000 mg ZnO-NPs kg− 1, MBC was reduced by 66%, and the bacterial community composition was considerably altered. Zhou and Zhang (2020) also showed that the microbial abundance was negatively influenced and bacterial structure was distorted by ZnO-NPs. The toxicity created by NPs may inhibit or even kill sensitive soil microorganisms, leading to a decline in soil respiration. However, the remaining resistant community contains taxa with a higher respiratory energy demand for maintenance, which means a greater loss of CO2 per unit of microbial biomass. Thus, the enhanced respiratory activity of the remaining resistant microbes is functionally compensatory to the suppressed activity of the taxa in the original community (Ge et al., 2011).
3.6. Effects of ZnO-NPs on soil microbial metabolic quotient
Microbial metabolic quotient (qCO2) is defined as the respiratory CO2-C evolved per unit of microbial biomass, which depends on both the abundance and activity of the soil microbial community. It is widely viewed that qCO2 is raised when the soil microbial biomass is allocating a larger proportion of C to maintenance than biomass synthesis (Anderson and Domsch, 1985). In the current study, the addition of ZnO-NPs induced a significant reduction in soil microbial biomass but had a smaller impact on soil respiration. Hence the qCO2 significantly increased in all ZnO-NPs concentrations applied (Fig. 5).
In the AH-amended soil samples, the qCO2 increased significantly (P < 0.05) with an increasing concentration of ZnO-NPs as it was the lowest (57.00 mg CO2-C g− 1 MBC d− 1) in uncontaminated soil, and highest (96.40 mg CO2-C g− 1 MBC d− 1) in soil contaminated with 1000 mg ZnO-NPs kg− 1 (Fig. 5). In WS-amended soil, the qCO2 was 71.0 mg CO2-C g− 1 MBC d− 1 in the control soil sample and had no statistical differences with those of the soil samples containing 100 to 600 mg ZnO-NPs kg− 1. However, the higher ZnO-NPs concentrations, i.e., 800 and 1000 mg ZnO-NPs kg− 1, caused significant (P < 0.05) increases in the qCO2 value in WS-amended soil (Fig. 5). The results, therefore, demonstrated that qCO2 increased remarkably with increasing ZnO-NPs concentration (Fig. 5), suggesting that biomass reduction in ZnO-NPs contaminated soils is largely due to the inefficient biosynthesis. These findings are in line with that reported by Rashid et al. (2017), showing that contamination of a leaf litter-amended soil with 1000 mg ZnO-NPs kg− 1 induced a significantly higher qCO2 as compared to the corresponding uncontaminated soil. Huang et al. (2022) also found that the addition of 150 and 300 mg ZnO-NPs kg− 1 increased the qCO2 in soil. The increase in qCO2 may also indicate a change in microbial community composition induced by ZnO-NPs, as several previous studies have shown ZnO-NPs can alter soil microbial community structures (Collins et al., 2012; Hänsch and Emmerling, 2010; Shen et al., 2015), which could easily cause changes in the respiration patterns.