*3.2. Biochar E*ff*ects on SOM, pH and Water Content*

Changes in soil properties, as a function of biochar application rate in the long-and short-term experiments, are shown in Figure 2. The SOM content significantly increased by 8.1–38.5% (long-term) and 3.9–49.3% (short-term), related to increasing biochar application rate (Figure 2A). The long- and short-term SOM changes remained statistically unchanged over time. Others have also noted similar SOM responses due to biochar application. For example, orchard prunings biochar (10%, *v:v*) reduced exchangeable metal concentrations in soil, which in part was attributed to increasing SOM [40].

Soil water contents were also significantly affected by increasing biochar application rates (Figure 2B). Soil water content increased by 3.8–41.6% (long-term, except 2016) and 4.5–77.4% (short-term), suggesting that biochar application may have improved soil physical properties. Positive changes in soil water content via biochar application have been observed by others [41–44].

Increasing biochar application rates also significantly increased soil pH, with soil pH having been shown to directly influence heavy metal fractions [45]. Soil pH significantly increased by 0.05–0.31 pH units (long-term) and 0.07–0.35 pH units (short-term), with greater changes associated with greater biochar application rates. The short-term usually had a greater effect on increasing soil pH compared to the long-term under the same treatment (Figure 2C). Increasing soil pH has been found to be a key factor for reducing heavy metal bioavailability [15,46].

**Figure 2.** The effect of long- (L) and short-term (S), increasing biochar application rates [0(C0), 10 (C1), 20 (C2) and 40 (C3) t ha<sup>−</sup>1] on soil (**A**) SOM, (**B**) water content, and (**C**) pH at rice harvest. Different lower-case letters above error bars indicate significant differences between the biochar treatments in either the short- or long-term for a given year [*p* < 0.05, Least significant difference (LSD) post-hoc test]. Error bars represent standard deviation of the mean (*n* = 3).

### *3.3. Biochar E*ff*ects on Cd and Pb Transfer in Rice*

The Cd and Pb concentrations in rice, as a function of biochar application rate in the long- and short-term, are shown in Figure 3. Following either short- or long-term biochar application at 40 t ha<sup>−</sup>1, rice Cd and Pb concentrations decreased by 30.7% and 45.2% (2016), 45.0% and 40.0% (2017), and 84.1% and 28.1% (2018), respectively, as compared to the control; lesser reductions were observed with lower biochar application rates. The rice husk, stem and root Cd and Pb concentrations followed similar trends. Biochar significantly reduced Cd and Pb transfer from roots to stems to grain; root Cd and Pb concentrations were over 40 mg kg−<sup>1</sup> and 300 mg kg<sup>−</sup>1, respectively, yet rice grain Cd and Pb concentrations were approaching or met national food safety level (≤0.2 mg kg<sup>−</sup>1) [47]. Other studies have shown that brinjal (i.e., eggplant) fruit Cd concentrations can be significantly decreased (up to 86.6%) using miscanthus biochar (1.5%, *w:w*) [48].

The BCF and TF were used to assess Cd and Pb transfer from the soil to the above-ground plant, and from plant roots to above-ground tissues, respectively. The BCF decreased by up to 77.1% (Cd) and 33.2% (Pb) in the long-term, and by up to 45.7% (Cd) and 42.3% (Pb) in the short-term. The TF also decreased by up to 53.2% (Cd) and 17.1% (Pb) in the long-term, and by up to 21.9% (Cd) and 23.2% (Pb) in the short-term (Table S2). The trend of decreasing BCF and TF with increasing biochar application rate supports our previous findings that biochar can reduce Cd and Pb bioavailability, suggesting that biochar may play a role in decreasing Cd and Pb transfer in the food chain [49]. Similar to our study, Mujtaba Munir et al., found that the bamboo biochar treatments reduced TF Cd and Pb by 49.6–61.0% and 61.0–70.7%, respectively, as compared to a control [21]; findings suggested that biochar effectively reduced bioavailable metal phases, leading to reduced metal translocation within plants.

**Figure 3.** The effect of long- and short-term, increasing biochar application rates [0 (C0), 10 (C1), 20 (C2) and 40 (C3) t ha<sup>−</sup>1] on Cd and Pb distribution in rice over time ((**A**): 2016 Cd; (**B**): 2016 Pb; (**C**): 2017 Cd; (**D**): 2017 Pb; (**E**): 2018 Cd; (**F**): 2018 Pb; Long: long-term; Short: short-term). Error bars represent standard deviation of the mean (*n* = 3).

## *3.4. Biochar E*ff*ects on Soil Enzyme Activity and Microbial Diversity*

Soil enzyme activities are key indicators of ecological change, and are particularly sensitive to anthropogenic modifications of heavy metal contaminated soils. Enzymatic activity alterations, due to soil amendment applications, are directly expressed in strength of biochemical reactions and their associated implications within soils [50]. In particular, soil oxidoreductase enzymatic activities are sensitive to change, and include alkaline phosphatase, dehydrogenase, urease and sucrase enzymatic activities. Specifically, Tabatabai suggest that alkaline phosphatase describes a broad group of enzymes that catalyze the hydrolysis of both ester and anhydride organic P, leading to increased inorganic P availability [51]. Dehydrogenase activity is typically considered a measure of general microbial activity [52]. Urease activity has long been known to hydrolyze urea to ammonium [53], thus increasing N availability to plants. And, sucrase activity is involved in sucrose degradation as well as direct SOM metabolism, which would enhance nutrient availability [54].

Soil enzymatic activities, as a function of biochar application rate in the long- and short-term, are shown in Figure 4. Biochar applications led to increased enzymatic activities in this Cd and Pb contaminated soil. Increasing biochar application rate increased: (a) phosphatase enzyme activity by 8.8–104.9% (long-term) and 9.0–83.6% (short-term); (b) dehydrogenase enzyme activity by 9.3–118.5% (long-term) and 10.4–103.5% (short-term); and (c) urease and sucrase enzyme activity by 91.3% and 77.9% (long-term), 132.4% and 83.6% (short-term), respectively. One could construe increased enzymatic activities as indicators of positive ecosystem change.

**Figure 4.** The effect of long- and short-term, increasing biochar application rates [0 (C0), 10 (C1), 20 (C2) and 40 (C3) t ha<sup>−</sup>1] on (**A**) alkaline phosphatase, (**B**) dehydrogenase, (**C**) urease, and (**D**) sucrase activity. Different lower-case letters above error bars indicate significant differences between the biochar treatments in either the short- or long-term for a given year (*p* < 0.05, LSD post-hoc test). The error bars represent standard deviation of the mean (*n* = 3).

Others have also noted enzymatic activity changes associated with biochar application. Chen et al., showed that wood biochar (4%, *w:w*) increased sucrase activity by up to 12.5-fold as compared to a control, with a subsequent improvement in soil fertility status [55]. Biochar amendment caused both catalase and urease activities to gradually decrease up to 45 days, but then increased over longer timeframes [18]. However, others biochar studies have found opposite responses. Liu et al., reported decreased urease activity in biochar amended soils, likely the result of oxidative reactions with free radicals on biochar surfaces [56]. Huang et al., utilized rice straw biochar (5%, *w:w*), observing a decrease in urease and alkaline phosphatase activities [57]. Regardless, in our study, biochar applications appear to cause positive changes in soil enzymatic activity, and thus biochar use may lead to positive changes in ecosystem functionality.

Increasing biochar application rates tended to increase the Shannon microbial diversity index in both the long- and short-term (Figure 5A). The Shannon index increased by 8.2% and 3.4% (2016), and by 46.2% and 29.8% (2017) in the long- and short-term studies, with greatest changes associated with the highest biochar application rate. Similar Shannon diversity index changes associate with bacterial communities in biochar amended soils have been previously reported [58]. Zhang et al., found that biochar application (1.5%) had the highest richness estimators and Shannon diversity index in amended, Cd contaminated soil [59].

**Figure 5.** The effect of long- and short-term, increasing biochar application rates [0 (C0), 10 (C1), 20 (C2) and 40 (C3) t ha<sup>−</sup>1] on changes of the (**A**) Shannon index and soil microbial diversity ((**B**): 2016; (**C**): 2017; (**D**): 2018). Different lower-case letters above error bars indicate significant differences between the biochar treatments in either the short- or long-term for a given year (*p* < 0.05, LSD post-hoc test). The error bars represent standard deviation of the mean (*n* = 3).

High-throughput PCR screening was used to further distinguish changes in microbial diversity (Figure 5B–D). During the studied years, microbial diversity increased by over 90% compared at phylum level. Steinbeiss et al., found similar responses due to biochar application. In the current study, biochar had the greatest effect on nine-phylum level microbial diversities in the short-term, significantly increasing them by 9.4–102.9% compared to the control (except *Chloroflexi* in the short-term in 2016) [60]. The nine-phylum level microbial diversities were slightly decreased by 6.4–46.9% in the long-term, except *Chloroflexi, Acidobacteria,* Unassigned and *Gemmatimonadetes*. Biochar (2%, *w:w*) has been shown to provide a positive effect on bacteria and invertebrate (such as earthworm) growth in heavy metal-contaminated agricultural soil in the short-term [2]. In a lab experiment, bacterial counts were increased by 149.4% compared with a control after 63 days following coconut shell biochar (5%, *w:w*) application to a Cd and Zn contaminated soil [38]. As with previous studies, in the current study, the soil microbial diversity was more greatly influenced in the short- versus long-term.
