The Fate of Heavy Metals and Risk Assessment of Heavy Metal in Pyrolysis Coupling with Acid Washing Treatment for Sewage Sludge

Pyrolysis is an emerging and effective means for sludge disposal. Biochar derived from sludge has broad application prospects, however, is limited by heavy metals. In this study, the fate of heavy metals (HMs) in pyrolysis coupling with acid washing treatment for sewage sludge was comprehensively investigated for the first time. Most of the HMs redistributed in the pyrolyzed residues (biochar) after pyrolysis, and the enrichment order of the HMs was: Zn > Cu > Ni > Cr. Compared with various washing agents, phosphoric acid presented a superior washing effect on most heavy metals (Cu, Zn, and Cr) in biochars derived at low pyrolysis temperature and Ni in biochars derived at high pyrolysis temperature. The optimal washing conditions for heavy metals (including Cu, Zn, Cr, and Ni) removal by H3PO4 were obtained by batch washing experiments and the response surface methodology (RSM). The total maximum HM removal efficiency was 95.05% under the optimal washing specifications by H3PO4 (acid concentration of 2.47 mol/L, L/S of 9.85 mL/g, and a washing temperature of 71.18 °C). Kinetic results indicated that the washing process of heavy metals in sludge and biochars was controlled by a mixture of diffusion and surface chemical reactions. After phosphoric acid washing, the leaching concentrations of HMs in the solid residue were further reduced compared with that of biochar, which were below the USEPA limit value (5 mg/L). The solid residue after pyrolysis coupling with acid washing resulted in a low environmental risk for resource utilization (the values of the potential ecological risk index were lower than 20). This work provides an environmentally friendly alternative of pyrolysis coupling with acid washing treatment for sewage sludge from the viewpoint of the utilization of solid waste.


Introduction
The amount and annual generation of sewage sludge (SS) have skyrocketed with the rapid development of the global urbanization and industrialization in recent years. In China, the amount of SS exceeded 50 million tons in 2020 and may reach 80 million tons by 2025 [1]. SS has been regarded as a bio-resource due to its contents of organic matter and nutrients (N, P, K). Meanwhile, there are numerous forms of harmful organic matter (e.g., pathogens, bacteria carrying antibiotic resistance genes (ARGs) [2], hazardous organic micro-pollutants, aromatic amines and polycyclic aromatic hydrocarbons (PAH)) and high levels of heavy metals (HMs) in SS, which have a huge potential risk to the environment [3,4]. Therefore, it has become an important global concern to develop an environmentally friendly approach for the disposal and utilization of SS. fate of HMs in sludge in the process of pyrolysis coupled with acid washing and facilitate the sludge-derived biochar application in the field of waste management.

Materials and Reagents
SS sample which had been dehydrated by a hydrothermal treatment coupling plateframe pressure filtration was collected from a wastewater treatment plant in Xiamen, China. The sludge was in the shape of a hard cake and the general characteristics of SS were investigated, shown in Table 1. The sludge was dried in a drying oven at 105 • C for 24 h until constant weight, and then sieved to a particle size of <125 µm. The reagents used in the experiment are supplied by Sinopharm Chemical Reagent. All these reagents are of analytical purity.

Pyrolysis Experiments
The pyrolysis experiment was conducted in a fixed-bed apparatus under nitrogen flow. The steps are similar to the previous experimental study [1]. According to the pyrolysis temperature x, the sludge-based biochar was named BC-X (Temperature).

Analytical Methods
Thermogravimetric (TG) analysis was performed using a TG analyzer (Netzsch TG 209, Selb, Germany), and the sludge sample was heated from room temperature to 1000 • C under nitrogen flow and a heating rate set as 10 • C/min. The contents of C, H, N, S and O elements of the sample were determined by an elemental analyzer (Vario MAX, Hanau, Germany). To determine the heavy metals in SS and BCs obtained from pyrolysis, the mixed acids (HNO 3 :HF:HClO 4 = 5:5:2, v/v, total 12 mL) in three replicates were injected into the PTFE digestion tube. The specific information about the digestion program steps is in the supplement material. The total concentration of the HMs was determined by the detection of inductively coupled plasma mass spectrometry (ICP-MS) (Agilent 7500 CX, Santa Clara, CA, USA) for the filtrate after the digestion was completed. Error bars represent standard deviations (n = 3).

Acid Washing Experiments
To explore the influence of different washing agents on HMs extraction in the SS and BCs, the samples were washed by various washing agents for 4 h. Washing experiments were conducted in the conical flasks. The vessels were shaken at 200 r/min in a constant temperature shaker. Typically, a weighed amount of SS and BCs was dispersed into 80 mL of washing solution and the range of liquid to solid ratios was from 2 to 12 mL/g. The washing experiments were performed under different acid concentrations (0.05, 0.2, 1.0, 5.0 M), different contact times (from 10 min to 24 h) and three different washing temperatures (25,50 and 75 • C) in order to determine the effect of different washing conditions for the HMs concentrations in the leachate. After washing, the mixture was centrifuged at 10,000 rpm for 10 min and filtered using a 0.22 µm hydrophilic filter. The HM content of the leachate was determined by ICP-OES. The collected solid was washed three times with deionized water and then dried at 85 • C for 24 h, collected for subsequent determination. All the tests were performed in triplicate and the average values and margin of error were reported for data analysis. The results were presented in this study as the average value and the error was calculated based on the standard deviation formula in Excel.

Optimization Procedure Experimental Design
Based on Box-Behnken design (BBD), the HMs removal washing process was further optimized by response surface methodology (RSM). The level and range of the essential factors (H 3 PO 4 concentration, washing temperature, and L/S ratio) influencing the washing efficiency were shown in Table S1. In the optimization experiments, the total removal rate of HMs was considered as the response variable. Design Expert software (Version 8.0.5) was introduced to fit the relationship between the efficiency of HM removal and the washing parameter. A quadratic equation model was used to predict the optimal value and elucidate the interaction between the washing parameter as Equation (1): where Y is the theoretical value; A 0 is a constant; α i , α ii , and α ij are the model's linear, quadratic, and interactive coefficients, respectively; n is the number of variables; and X i and X j are the coded independent variables.

Leachability (TCLP) and Distribution of Heavy Metals (BCR)
The leachability of HMs in the sample was determined by the toxicity characterization (TCLP, US EPA Method 1311) [28]. The chemical speciation of the HMs samples before and after the washing treatment were determined by BCR (Bureau Community of Reference). Each of the four fractions was defined as: F1(exchangeable fraction), F2 (reducible fraction), F3 (oxidizable fraction) and F4 (residual fraction). The specific steps were described in the previous study [29].

Environmental Impact of Heavy Metals
The potential ecological risk index (RI) was usually used to assess the environmental impact of the HMs, which has been adopted by most researchers [5]. The actual calculation program uses the following equations: where C f represents the contamination factor of an individual HM; C 1 , C 2 , C 3 and C 4 are the concentrations of the four fractions of the HMs, respectively; E r is the potential ecological risk factor for each HM; and Tr values were the toxic factor of each HM, for Cr, Ni, Cu, Zn, Cd, and Pb were 2, 6, 5, 1, 30, and 5, respectively [5]. The RI values identified four types of risk levels, where RI ≤ 50 represents "low" risk, 50 < RI ≤ 100 represents "moderate" risk, 100 < RI ≤ 200 represents "considerable" risk, and RI > 200 represents "high" risk.

Properties of SS and BCs
The general characteristics of SS and its biochars at different temperatures were investigated, shown in Table 1. The moisture, ash, volatile and fixed carbon contents in SS were 40.28%, 63.74%, 35.07% and 1.19%, respectively. The contents of C, H, N, S and O were 15.44%, 3.25%, 1.62%, 0.55% and 15.4%, respectively. With the increase of pyrolysis temperature, the yield of biochar decreased from 74.64% of the original feedstock mass at 400 • C, 69.86% at 600 • C, to 61.16% at 800 • C. The volatile matter decreased from 35.07% for the raw sludge, 14.65% for the biochar prepared at 400 • C to 1.6% at 800 • C. The ash concentration in the biochar significantly increased while the fixed carbon of all biochar was slightly reduced, which was consistent with other researchers [5,30]. The decrease of the C, H and N content in all the biochar samples was observed from 400 to 800 • C, which was due to the increased volatilization containing these elements during pyrolysis. In addition, The H/C ratio of BCs at different pyrolysis temperatures were all below 2.4, which showed that they were thermochemically converted materials that had a greater proportion of fused aromatic ring structures [18]. The BET surface area (SA) of the biochars was shown in Table 1 as is the increase in biochar surface area with the increasing pyrolysis temperature. For instance, the SA of the BC sample increased by 17.08 m 2 /g more than that of SS when pyrolysis temperature reached 400 • C, whereas the SA of BC800 increased to 49.68 m 2 /g.
To explore the influence of temperature on the thermal decomposition of the sludge, TG analysis was performed in a N 2 atmosphere. The mass loss as a function of the pyrolysis temperature is depicted in Figure S2. The mass losses for sludge were divided into the following four stages: (1) the first stage was the volatilization stage from 105 to 480 • C, where the total weight loss of the sludge was 22.48%, and the weight loss reached the maximum weight loss rate at approximately 342.42 • C. The DTG peak was not singlepeaked, indicating that the pyrolysis in this stage included the superposition of a series of thermal decomposition reactions. This stage mainly included the volatilization and decomposition of fats, carbohydrates and other organic compounds in SS [29]. (2) The pyrolysis rate in the 480-640 • C range was significantly lower; the TG curve tended to be flat, and the weight loss was only 6.32%. It was thus inferred that the main pyrolysis reactions in the previous temperature range were complete. (3) The weight loss rate in this stage was relatively fast in the range of 640 to 825 • C, and the total weight loss in the decomposition stage of the intermediate product was 10.64%, where the maximum weight loss rate occurred at 800 • C from the DTG results, which was probably because the organic matter underwent decomposition, condensation, dehydrogenation, cyclization, etc., and simultaneously, some inorganic substances were thermally decomposed. (4) In the range of 825-1000 • C the weight loss was approximately 7.51%. According to DTG results, the maximum weight loss rate occurred at 925 • C. This stage was the decomposition stage for minerals, including carbonates, alkali metal oxides and chlorides [30]. The undecomposed substances were mainly ash and fixed carbon.

Total Concentration of Heavy Metals
Heavy metals have always been of concern in the disposal of sludge and the application of its biochar. Therefore, it is necessary to analyze the migration behavior of HMs in sludge pyrolysis. The total concentrations of HMs in SS were listed in Table S1. The HM contents in SS decreased as follows: Cr > Cu > Zn > Ni > Pb > As > Cd. Cr had the highest contents among the HMs in the SS, with a maximum content of 8195 mg/kg, which may be due to industrial grinding factories that may use a large number of chromiumcontaining reagents around this wastewater treatment plant (WWTP). Cu content was about 7945 mg/kg, due to the sewage pipe network being mainly copper pipe in China. According to the relevant Chinese legal standards, the content of Cu, Zn, Cr and Ni in SS exceeded the maximum permitted limit. Because the contents of As, Pb and Cd in SS and its biochars were far below the limit, the subsequent washing experiments focused on Cu, Zn, Cr and Ni in this study. In Figure 1a, the residual rates of Cu, Zn, Cr, and Ni in the biochars were over 96.41%, 98.74%, 94.58%, and 96.42%, respectively, which meant that most of heavy metals mainly redistributed in the biochars, which was consistent with the previous study [31]. Accordingly, shown in Figure 1b, the total concentrations of Cu, Zn, Cr and Ni in biochars increased due to the higher boiling temperatures of heavy metals, so the loss in weight of organic compounds during pyrolysis was higher than the loss in weight of heavy metals [3]. In Figure 1c, RE represented the relative enrichment factor of heavy metal elements, which was defined as Formula (5), where cbiochar and cfeed represented the concentrations of HMs in biochar and raw sludge, respectively. The larger the value of RE is than 1, the higher the enrichment degree of HMs in BCs is, and vice versa. Four kinds of HMs were all enriched at 400-800 °C, and the RE values were all above 1.20. Amongst this, the enrichment of Zn was the most obvious. According to the comprehensive comparison, the enrichment order of HMs in sludge biochar was Zn > Cu > Ni > Cr.

Speciation of Heavy Metals in the SS and Its Biochars
The bioavailability and ecotoxicity of HMs in SS and biochars are mainly associated with chemical speciation. The acid-extractable state (F1) and reducible state (F2), which are very prone to leaching, can directly enter plants through plant roots. The oxidizable (F3) fraction is related to the potentially bioavailable category. The residual state (F4) was unbioavailable because of its strong stability [28].
Shown in Figure 2, most of the Zn and Ni from SS were primarily distributed in the F1 + F2 fraction (>63%). In comparison, Cu and Cr presented lower environmental risks, as Cu was primarily distributed in the F3 + F4 fraction (73.64%) and Cr remained in the F3 + F4 fraction (>98%). This mainly related to the existence of organometallic or residual phases containing Cu and Cr in raw SS [32]. A remarkable decline of Cu and Zn in F1 and F2 occurred and a significant gradual increase to the F3 + F4 fraction of biochars with increasing pyrolysis temperature from 400-800 °C was presented simultaneously ( Figure  2). High pyrolysis temperature also led to the further transformation of Cr from F3 into F4, whereas Cr in BC600 remained in the F4 fraction (99%). The possible immobilization mechanisms were revealed as the decomposition into silicates or metal oxides, the transformation from the amorphous to crystalline state, and the embedding in the C matrix as organometallic compounds [28,33]. Nevertheless, the F3 + F4 fraction of Ni in biochars showed a trend of increasing following decreasing and the F3 + F4 fraction of Ni in BC400 reached the maximum (74.36%). This was probably due to the decomposition of the carbonate state (F1) and the organic bonded state (F3) of Ni in biochar in steps between 400 and 800 °C. In the sequential extraction, the recovery between the sum of the four fractions and the total HM concentration was between 98.65% and 102.47% (Table S2). As a quality control indicator, it showed satisfactory agreement. Although the pyrolysis process had made the HM immobilization to a certain extent, the potential environmental risk still existed due to the high concentrations of HMs which far exceeded the threshold values in the national standards. From the viewpoint of environmental safety to application, further HM washing should be necessary. In Figure 1c, RE represented the relative enrichment factor of heavy metal elements, which was defined as Formula (5), where c biochar and c feed represented the concentrations of HMs in biochar and raw sludge, respectively. The larger the value of RE is than 1, the higher the enrichment degree of HMs in BCs is, and vice versa. Four kinds of HMs were all enriched at 400-800 • C, and the RE values were all above 1.20. Amongst this, the enrichment of Zn was the most obvious. According to the comprehensive comparison, the enrichment order of HMs in sludge biochar was Zn > Cu > Ni > Cr.

Speciation of Heavy Metals in the SS and Its Biochars
The bioavailability and ecotoxicity of HMs in SS and biochars are mainly associated with chemical speciation. The acid-extractable state (F1) and reducible state (F2), which are very prone to leaching, can directly enter plants through plant roots. The oxidizable (F3) fraction is related to the potentially bioavailable category. The residual state (F4) was unbioavailable because of its strong stability [28].
Shown in Figure 2, most of the Zn and Ni from SS were primarily distributed in the F1 + F2 fraction (>63%). In comparison, Cu and Cr presented lower environmental risks, as Cu was primarily distributed in the F3 + F4 fraction (73.64%) and Cr remained in the F3 + F4 fraction (>98%). This mainly related to the existence of organometallic or residual phases containing Cu and Cr in raw SS [32]. A remarkable decline of Cu and Zn in F1 and F2 occurred and a significant gradual increase to the F3 + F4 fraction of biochars with increasing pyrolysis temperature from 400-800 • C was presented simultaneously (Figure 2). High pyrolysis temperature also led to the further transformation of Cr from F3 into F4, whereas Cr in BC600 remained in the F4 fraction (99%). The possible immobilization mechanisms were revealed as the decomposition into silicates or metal oxides, the transformation from the amorphous to crystalline state, and the embedding in the C matrix as organometallic compounds [28,33]. Nevertheless, the F3 + F4 fraction of Ni in biochars showed a trend of increasing following decreasing and the F3 + F4 fraction of Ni in BC400 reached the maximum (74.36%). This was probably due to the decomposition of the carbonate state (F1) and the organic bonded state (F3) of Ni in biochar in steps between 400 and 800 • C. In the sequential extraction, the recovery between the sum of the four fractions and the total HM concentration was between 98.65% and 102.47% (Table S2). As a quality control indicator, it showed satisfactory agreement. Although the pyrolysis process had made the HM immobilization to a certain extent, the potential environmental risk still existed due to the high concentrations of HMs which far exceeded the threshold values in the national standards. From the viewpoint of environmental safety to application, further HM washing should be necessary.

Optimal Washing Agent
To assess the efficiency of different washing agents on the efficiency of HM removal from the sludge-derived biochar, BC400 was extracted by washing agents (H2SO4, HNO3, H3PO4, HCl, citric acid and EDTA) of 1 M at a L/S of 10 mL/g for 4 h. The findings ( Figure  3a) suggested that the removal efficiency of HMs by inorganic acids was higher than organic acid, which were consistent with the report by [34]. In comparison, the removing percentages of Cu (98.26%), Zn (94.31%), Cr (9.11%) and Ni (52.26%) by H3PO4 were superior to those by the other inorganic acids. The mechanism of the phenomena may be that HCl, H2SO4 and HNO3 in solution can only provide H + and the possible reaction might be: The removal efficiency of HMs by H3PO4 were higher than those by H2SO4, HNO3, and HCl. Differing from the above mechanism, H3PO4, as a kind of ternary medium-strong acid, provides not only more H + ions, which will react with metal oxide, metal sulfide and

Optimal Washing Agent
To assess the efficiency of different washing agents on the efficiency of HM removal from the sludge-derived biochar, BC400 was extracted by washing agents (H 2 SO 4 , HNO 3 , H 3 PO 4 , HCl, citric acid and EDTA) of 1 M at a L/S of 10 mL/g for 4 h. The findings (Figure 3a) suggested that the removal efficiency of HMs by inorganic acids was higher than organic acid, which were consistent with the report by [34]. In comparison, the removing percentages of Cu (98.26%), Zn (94.31%), Cr (9.11%) and Ni (52.26%) by H 3 PO 4 were superior to those by the other inorganic acids. The mechanism of the phenomena may be that HCl, H 2 SO 4 and HNO 3

Optimal Washing Agent
To assess the efficiency of different washing agents on the efficiency of HM removal from the sludge-derived biochar, BC400 was extracted by washing agents (H2SO4, HNO3, H3PO4, HCl, citric acid and EDTA) of 1 M at a L/S of 10 mL/g for 4 h. The findings ( Figure  3a) suggested that the removal efficiency of HMs by inorganic acids was higher than organic acid, which were consistent with the report by [34]. In comparison, the removing percentages of Cu (98.26%), Zn (94.31%), Cr (9.11%) and Ni (52.26%) by H3PO4 were superior to those by the other inorganic acids. The mechanism of the phenomena may be that HCl, H2SO4 and HNO3 in solution can only provide H + and the possible reaction might be: H + + MexO/MexS/MexCO3 → xMe 2/x+ + H2O/H2S/CO2 The removal efficiency of HMs by H3PO4 were higher than those by H2SO4, HNO3, and HCl. Differing from the above mechanism, H3PO4, as a kind of ternary medium-strong acid, provides not only more H + ions, which will react with metal oxide, metal sulfide and The removal efficiency of HMs by H 3 PO 4 were higher than those by H 2 SO 4 , HNO 3 , and HCl. Differing from the above mechanism, H 3 PO 4 , as a kind of ternary medium-strong acid, provides not only more H + ions, which will react with metal oxide, metal sulfide and metal carbonate, etc. Meanwhile, the PO 4 3− ions, which were produced H + being ionized, also had a larger complexing ability to metal ions [35]. Therefore, the remove efficiency by H 3 PO 4 is higher than other inorganic acids under the same conditions. These multiple synergisms of H 3 PO 4 constituted the best removal efficiency results. The possible reactions might be as below: H + + PO 4  The removal efficiency of HMs relates not only to the chemistry of the extractant, as well as the sample geochemistry and characteristics of the metals, but also on dosage of extractants and the process conditions [36]. In Figure 3b, When the L/S ratio was 2 mL/g, the Cu, Zn, Cr and Ni removal efficiencies during H 3 PO 4 washing were 68.27%, 43.26%, 25.27%, and 28.10%, while the L/S ratio increased to 8 mL/g, the values reached at 98.01%, 84.68%, 80.44%, and 47.44%, respectively. Consequently, the HM removal efficiencies increased with a higher L/S ratio.

Effect of H 3 PO 4 Concentration on the Removal Efficiency of HMs
The removal efficiencies of Cu, Zn, Cr and Ni in SS and biochars varied with the concentration of H 3 PO 4 ( Figure 4). The removal efficiencies of HMs in SS and biochars significantly increased with the increase of washing agent concentration from 0.05 M to 5 M, which was similar to that reported previously [22].  4 were similar with that of sludge. The effect of pyrolysis temperature for the removal efficiency of HMs was different. Different pyrolysis temperatures have little effect on the removal efficiency of Cu. When the H 3 PO 4 concentration was 0.2 M, the Cu removal efficiencies of SS, BC400, BC600 were 96.51%, 98.67%, and 99.25%, respectively. Additionally, the removal efficiency of Cu in BC800, decreasing slightly, was about 94.53%. However, the removal efficiency of Zn and Cr in biochars were affected by the pyrolysis temperature greatly. When the H 3 PO 4 concentration was 5 M, the removal efficiencies of Zn in BC400, BC600 and BC800 were 94.52%, 50.64% and 8.49%, while these of Cr were 96.77%, 53.45%, and 38.89%, respectively. On the contrary, the removal efficiency of Ni in biochars increased from 400 to 800 • C. The removal value of Ni in BC400, BC600, BC800 were 52.57%, 71.68%, and 97.70%, respectively, which corresponded to the BCR result. The above results are mainly related to the different speciation of HMs during sludge pyrolysis.

Effect of Washing Temperature on the Removal Efficiency of Heavy Metals
The effect of washing temperature on the removal efficiencies of HMs in SS and BCs was shown in Figure 5. The removal efficiency of Cr and Zn in sludge and biochars gradually increased with the increase of washing temperature. The washing effect of Zn and Cr at 75 °C increased significantly compared with that at 25 and 50 °C, especially for BC600. The removal efficiencies of Zn and Cr in BC600 increased from 69.90% and 29.55% to 93.21% and 88.62%, respectively, when the leaching temperature increased from 50 °C to 75 °C. Moreover, the removal efficiency of Cu and Ni in biochar with higher pyrolysis temperatures decreased significantly when the washing temperature was 75 °C. Considering the washing effect of various heavy metals, the optimal leaching temperature is about 50 °C.

Effect of Washing Temperature on the Removal Efficiency of Heavy Metals
The effect of washing temperature on the removal efficiencies of HMs in SS and BCs was shown in Figure 5. The removal efficiency of Cr and Zn in sludge and biochars gradually increased with the increase of washing temperature. The washing effect of Zn and Cr at 75 • C increased significantly compared with that at 25 and 50 • C, especially for BC600. The removal efficiencies of Zn and Cr in BC600 increased from 69.90% and 29.55% to 93.21% and 88.62%, respectively, when the leaching temperature increased from 50 • C to 75 • C. Moreover, the removal efficiency of Cu and Ni in biochar with higher pyrolysis temperatures decreased significantly when the washing temperature was 75 • C. Considering the washing effect of various heavy metals, the optimal leaching temperature is about 50 • C.

Effect of Washing Time and the Kinetic Study
Regarding the effect of washing time on the removal of heavy metals in the SS and biochars using 1 M H3PO4 and a L/S ratio of 1:1 as shown in Figure 6, the concentrations

Effect of Washing Time and the Kinetic Study
Regarding the effect of washing time on the removal of heavy metals in the SS and biochars using 1 M H 3 PO 4 and a L/S ratio of 1:1 as shown in Figure 6, the concentrations of HMs in the washing effluent increased with time. The removal efficiency of most HMs, except Ni in BC600, presented a rapidly increasing trend then stood in a stable stage, which is consistent with previous studies about HMs leaching from sludge ash [32]. The fundamental molecular theory of liquids was used to explain this result, in other words, the initial acid has the largest hydrogen ion concentration lead to the fastest reaction speed. Subsequently, the hydrogen ion concentration and activity of the acids gradually drop with increasing washing time [37]. The removal efficiency of heavy metals nearly reached equilibrium at about 4 h. A kinetic study on the removing characteristics of HMs in SS and BCs was performed. The washing process is a heterogeneous reaction process in a liquid-solid reaction system [38]. The shrinking core model (SCM) is the most widely used and is considered to be the kinetic model that best reflects the actual situation of the leaching process [39]. Herein, assuming the residue particle is spherical, the H3PO4-residue reaction process could be described by the shrinking core model, and was divided into two independent stages, which are controlled by chemical reaction as Equation (6), diffusion of the reagent or product layer as Equation (7), respectively. If the diffusion or the surface chemical reactions are the slowest steps, the equations of the shrinking core models are expressed as follows, respectively: However, the R 2 values of fitting degrees are below 90%, relatively low when using the above equation alone. In order to characterize the kinetic process of the solid-liquid heterogeneous reaction more accurately, based on the shrinking nucleation model of spherical solid particles, segmental fitting of different reaction times was carried out. Table  2 listed the mixed control kinetic model. It is found that R 2 is greater than 0.95, indicating that the removal of HMs from sludge and biochars is controlled by a mixture of diffusion and surface chemical reactions.   A kinetic study on the removing characteristics of HMs in SS and BCs was performed. The washing process is a heterogeneous reaction process in a liquid-solid reaction system [38]. The shrinking core model (SCM) is the most widely used and is considered to be the kinetic model that best reflects the actual situation of the leaching process [39]. Herein, assuming the residue particle is spherical, the H 3 PO 4 -residue reaction process could be described by the shrinking core model, and was divided into two independent stages, which are controlled by chemical reaction as Equation (6), diffusion of the reagent or product layer as Equation (7), respectively. If the diffusion or the surface chemical reactions are the slowest steps, the equations of the shrinking core models are expressed as follows, respectively: However, the R 2 values of fitting degrees are below 90%, relatively low when using the above equation alone. In order to characterize the kinetic process of the solid-liquid heterogeneous reaction more accurately, based on the shrinking nucleation model of spherical solid particles, segmental fitting of different reaction times was carried out. Table 2 listed the mixed control kinetic model. It is found that R 2 is greater than 0.95, indicating that the removal of HMs from sludge and biochars is controlled by a mixture of diffusion and surface chemical reactions.

Optimization of the removal of Heavy Metals
The combined effect of washing parameters (including acid concentration, washing temperature, and L/S ratio) on the removal of HMs was further determined by response surface methodology (RSM). The range for the washing specifications were determined from the above single variable experimental results. 17 runs of the total were performed to optimize three variables the values of which are shown in Table S3. A quadratic equation model represented the removal efficiencies for HMs by H 3 PO 4 as follows: Y = 1.18X 1 + 1.69X 2 + 54.56X 3 − 0.03X 1 X 2 − 2.04X 1 X 3 − 0.01X 2 X 3 + 9.87X 1 2 − 0.015X 2 2 − 2.86X 3 2 − 196.09, where Y represents the HMs removal efficiency; X 1 , X 2 , and X 3 represent the acid concentration, washing temperature, and washing time, respectively.
The extent of the equation fitting was evaluated by ANOVA analysis (Table S4). The F-value was 15.77, which meant that the second-order polynomial equation models was adequate for the optimization study. The p-value represents the significance of the influential washing factors and reflect their interaction strength [37]. In this work, the p-value of X 1 was much less than 0.05, which meant that the H 3 PO 4 concentration(C 0 ) was significant in the removal process of HMs. Moreover, the smaller X 1 and X 2 p-value (p = 0.0032) indicated a proper relationship between C 0 and L/S ratio. In this study, R 2 (0.9882) and R 2 (adj.) (0.9768) were close to 1.0, which meant that the significance and agreement between the experimental and predicted values were great. Furthermore, the precision and reliability of the developed models was high because the coefficient of variation value (2.01) was relatively low [40]. In order to evaluate the combined effects of initial concentration, washing temperature and L/S ratio on the removal efficiency of washing by H 3 PO 4 , the 3D mesh diagrams were plotted as shown in Figure 7. According to the response surface analysis, the optimal washing specifications for Cu, Zn, Cr, and Ni removal by H 3 PO 4 were below: acid concentration of 2.47 M, L/S of 9.85 mL/g, and washing temperature of 71.18 • C. Under these specifications, the predicted maximum HM removal efficiency reached up to 95.61%. Verification experiments were conducted in triplicate at the optimal washing specifications to validate the model suitability. The total HMs removal efficiency was 95.05%, which was in accordance with the predicted values, and the result indicated that the RSM model was accurate for optimizing washing experiment.

Chemical Speciation of Heavy Metals and Environmental Risk Evaluation of SS and B after Washing
The bioavailability and ecotoxicity of HMs in SS and biochars after the washin cess are associated with both of the concentration and chemical speciation [30]. Th centration and speciation distribution of HMs in SS and biochars were shown in Fig  Compared with the initial HMs distribution in the SS and biochars, the concentrati HMs in SS and biochars after the washing process were reduced significantly. Th centrations of Cu in SS and BC400 after washing were 73.01 mg/kg and 190.25 m respectively, which were both below the control standards for agriculture use (Le 500 mg/kg). These in BC600 and BC800 after washing were 645.10 mg/kg and 1 mg/kg, respectively, both below the control standards for agriculture use (Level B mg/kg). The concentrations of Zn in SS and BC400 after washing were 30.03 mg/k 339.99 mg/kg, respectively, which were below the control standards for agricultu (Level A, 1500 mg/kg). Furthermore, the one in BC600 after washing was slightly the control standards for agriculture use (Level B, 3000 mg/kg), about 2980.76 mg/ for the concentrations of Cr in SS and BC400, they were 114.97 mg/kg and 908.12 m respectively after washing, which were both below the control standards for agric use (Level B, 1000 mg/kg). Moreover, the concentrations of Ni in SS and BC800 after ing were below the control standards for agriculture use (Level B, 200 mg/kg), wh concentrations of Cr in BC800 was 43.99 mg/kg, far below the above standards.

Chemical Speciation of Heavy Metals and Environmental Risk Evaluation of SS and BCs after Washing
The bioavailability and ecotoxicity of HMs in SS and biochars after the washing process are associated with both of the concentration and chemical speciation [30]. The concentration and speciation distribution of HMs in SS and biochars were shown in Figure 8. Compared with the initial HMs distribution in the SS and biochars, the concentrations of HMs in SS and biochars after the washing process were reduced significantly. The concentrations of Cu in SS and BC400 after washing were 73.01 mg/kg and 190.25 mg/kg, respectively, which were both below the control standards for agriculture use (Level A, 500 mg/kg). These in BC600 and BC800 after washing were 645.10 mg/kg and 1450.17 mg/kg, respectively, both below the control standards for agriculture use (Level B, 1500 mg/kg). The concentrations of Zn in SS and BC400 after washing were 30.03 mg/kg and 339.99 mg/kg, respectively, which were below the control standards for agriculture use (Level A, 1500 mg/kg). Furthermore, the one in BC600 after washing was slightly below the control standards for agriculture use (Level B, 3000 mg/kg), about 2980.76 mg/kg. As for the concentrations of Cr in SS and BC400, they were 114.97 mg/kg and 908.12 mg/kg, respectively after washing, which were both below the control standards for agriculture use (Level B, 1000 mg/kg). Moreover, the concentrations of Ni in SS and BC800 after washing were below the control standards for agriculture use (Level B, 200 mg/kg), while the concentrations of Cr in BC800 was 43.99 mg/kg, far below the above standards. Compared with the HM distribution before washing, the washing process prom significantly the reduction of the F1 of Cu, Zn and Ni in SS and BCs especially the SS and BC400, which declined by 91.38% and 68.06%. The F2 fraction Zn in SS, BC60 BC800 decreased by 96.23%, 94.58% and 97.02%, while that of Ni in BC800 decline 96.82%. The proportions of F1 and F2 of Cr had little change after washing, less tha In the biochars (BC600, BC800) obtained at higher pyrolysis temperatures, the propo of F3 decreased while that of F4 increased. The F4 proportions of all four metals in sl and biochars increased, shown in Figure 6. The loss of more easily extractable frac and lattice disruption in solid residues was the main reason for the F4 fraction inc among the HMs after washing [41].
To further quantify the environmental effect and potential environmental ri sludge and BCs after washing (AWSS and AWBC) with H3PO4, the potential ecolo risk index (RI) values of AWSS and AWBCs were assessed in this work. The RI value is about 67.21 (between 50-100), shown in Table 3, indicating a "moderate" risk t environment. The RI values of BCs, except BC800, were all between 50-100 and wer atively decreased than that of SS, indicating "moderate" risk [29]. However, becaus total concentrations of HMs and BCs were large, the leaching risk was still high. Aft washing of the solid residue, the RI value decreased significantly to below 50, mea that the degree of environmental risk was "low".  Compared with the HM distribution before washing, the washing process promoted significantly the reduction of the F1 of Cu, Zn and Ni in SS and BCs especially the Zn in SS and BC400, which declined by 91.38% and 68.06%. The F2 fraction Zn in SS, BC600 and BC800 decreased by 96.23%, 94.58% and 97.02%, while that of Ni in BC800 declined by 96.82%. The proportions of F1 and F2 of Cr had little change after washing, less than 1%. In the biochars (BC600, BC800) obtained at higher pyrolysis temperatures, the proportion of F3 decreased while that of F4 increased. The F4 proportions of all four metals in sludge and biochars increased, shown in Figure 6. The loss of more easily extractable fractions and lattice disruption in solid residues was the main reason for the F4 fraction increase among the HMs after washing [41].
To further quantify the environmental effect and potential environmental risk of sludge and BCs after washing (AWSS and AWBC) with H 3 PO 4 , the potential ecological risk index (RI) values of AWSS and AWBCs were assessed in this work. The RI value of SS is about 67.21 (between 50-100), shown in Table 3, indicating a "moderate" risk to the environment. The RI values of BCs, except BC800, were all between 50-100 and were relatively decreased than that of SS, indicating "moderate" risk [29]. However, because the total concentrations of HMs and BCs were large, the leaching risk was still high. After the washing of the solid residue, the RI value decreased significantly to below 50, meaning that the degree of environmental risk was "low".  The degree of toxicity of HMs in the SS and its biochars is reflected in both the leaching content and the existing speciation of. The leaching concentration of HMs tested by TCLP are listed in Table S5. The leaching concentration of HMs in biochars was significantly lower than that in sludge samples, and the leaching concentration gradually reduced with the increasing pyrolysis temperature. However, HMs (except Cr) in sludge and biochar, especially in low-temperature pyrolytic produce (BC400), exceed the USEPA limits (5 mg/L). After H 3 PO 4 washing, leaching concentrations of HMs in solid residues were further reduced compared with that of biochar, and all were below the USEPA limit value. From the viewpoint of environmental safety, pyrolysis coupling washing treatment of sewage sludge should be necessary.

Conclusions
In this work, a novel process for SS treatment through pyrolysis coupled with acid washing was explored and the fate of heavy metals in tandem pyrolysis and washing treatment was comprehensively investigated for the first time. Most heavy metals (HMs) redistributed in the biochars after pyrolysis. The enrichment order of the HMs was: Zn > Cu > Ni > Cr. The phosphoric acid, among various washing agents, had a superior leaching effect on Cu, Zn, and Cr in biochars derived at a low pyrolysis temperature and Ni in biochars derived at a high pyrolysis temperature. According to the optimal results by the response surface methodology, the concentration was more significant for the removal of HMs and a stronger interaction strength between acid concentration and L/S ratio. Further, the total maximum HM removal efficiency reached 95.05%, which was in accordance with the predicted value by RSM, under the optimal conditions (acid concentration of 2.47 M, L/S of 9.85 mL/g, and washing temperature of 71.18 • C). Kinetic results indicated that the washing process was controlled by a mixture of diffusion and surface chemical reactions. After phosphoric acid washing, the leaching concentrations of HMs in the solid residue were further reduced significantly compared with that of biochar, which were below the USEPA limit value (5 mg/L). The solid residue after pyrolysis coupling washing had a low environmental risk for resource utilization (the values of RI were lower than 20). This work provides an alternative of pyrolysis coupling washing treatment for sewage sludge from the viewpoint of environmental safety. Further research would focus on the recovery of heavy metals based on the investigation about fate of HMs in sludge in the process and the evaluation of practical application of pyrolysis coupling washing treatment for sewage sludge.
Supplementary Materials: The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/toxics11050447/s1, Figure S1: Schematic diagram of the pyrolysis experimental apparatus; Figure S2: Thermogravimetric curve of chromium-rich tanning sludge; Table S1: Heavy metals concentration of sludge samples and relevant Chinese legal standards; Table S2: The recovery of the sum of four fractions ratio the total HMs concentration; Table S3: The levels of each variable and corresponding heavy metals removal from BC400 obtained from the Box-Behnken design; Table S4: ANOVA for Response Surface Quadratic Model; Table S5: The leaching concentration of heavy metals in sludge and biochars.